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SUBAQUEOUS DISPOSAL

OF REACTIVE MINE

WASTES: AN OVERVIEW

MEND Project 2.11.la

June 1989

Subaqueous Disposai of

Reactive Mine Wastes: An Overview

A British Columbia Acid Mine

Drainage Task Force Project

ln Contribution To MEND

(Mine Environment Neutral Drainage)

Prepared for:

British Columbia Ministry of Energy Mines and Petroleum Resources

Funded by:

The Canada/British Columbia Minerai Development Agreement

Prepared by:

Rescan Environmental Services Ltd. Vancouver, Canada

June 1989

(j(escan)

EXECUTIVE SUMMARY

Finding an environmentally safe, yet economical, method of disposing of reactive mine

wastes is a challenge facing both the mining industry and government. When such

materials contain sulphides, the conventional practice of land-based disposai has often

resulted in the generation of acidic water and the concomitant leaching of trace metals

from the mine wastes. Acid production in tailings and waste rock is a result of the

oxidation of sulphide minerais (principally iron pyrite). Acid generation results in

various, often severe, impacts on water chemistry and biological resources. The

environmental implications are considerable, particularly since the problem can persist

after active mining bas ceased.

One method for controlling acid generation which is receiving increasing attention is

the practice of depositing reactive mine wastes underwater. While the Metal Mining

Liquid Effluent Regulations, authorized under the federal Fisheries Act, currently

prohibit lake disposai of mine tailings, an exemption can be issued through a federal

cabinet Order-in-Council. An argument for subaqueous disposai is based on the

premise that acid generation is suppressed in submered mine wastes that are essentially

unexposed to oxygen and bacterial action. This suggestion is predicated on knowledge

of the biogeochemical nature of lacustrine sediments, and appears to be supported by

observations made in specific field studies. Hence, it is consistent to suggest that

sulphide-rich mine wastes may be disposed of underwater without significant release of

metals to the overlying waters. In some circumstances, however, a number of factors

can act individually or in concert to promulgate release of metals, with the associated

potential for environmental degradation. Nevertheless, with sufficient knowledge of

post-depositional chemical reactivity of the specific tailings and adherence to disposai

criteria, these factors can be mitigated to various extents and impacts on water quality

and indigenous biota minimized.

While this conventional wisdom supporting subaqueous disposai may be correct, only a

limited number of reviews and even fewer field studies have been undertaken that add

significant insight. Consequently, the efficacy of underwater disposai remains largely

unproven.

- 1 -

EXECUTIVE SUMMARY

Responding to the clear need for establîshing effective methods to mitigate acid mine

drainage, and recognizing the considerable promise subaqueous disposa! holds in this

regard, the British Columbia Acid Mine Drainage Task Force issued a call for

proposals to complete a literature review on underwater disposai of reactive mine

wastes. The study was to focus on the state-of-the-art in theoretical knowledge and case

studies of British Columbia mines disposing of wastes in a freshwater environment.

Consideration was to be given to potential impacts on freshwater biological systems,

including both physical and chemical effects of mine waste disposai. As a consequence,

Rescan Environmental Services Ltd. was retained to undertake the literature review on

the Subaqueous Disposa[ of Reactive Mine Wastes.

This study was completed based on a comprehensive rev1ew of all aspects of

subaqueous disposai of reactive mine wastes in a freshwater environment, both

theoretical and applied. While emphasis was placed on the British Columbia

experience with underwater disposal practices, as mandated in the RFP, attention was

also given to other parts of Canada and the United States. During the review,

numerous and varied sources of literature were consulted through both desk and

computer-aided search methods. Sources of information included, among others,

literature on acid mine drainage mechanisms and chemistry, aquatic chemistry,

geochemistry, hydrogeochemistry, biochemistry, microbiology, limnology and aquatic

biology/ecology. Whenever possible, relevant case study data were obtained which

often addressed one or more of the above areas to various degrees of detail. Although

the scope of this report is confined to freshwater (i.e. lake disposai), the literature

review includes documentation on land-based tailings and leach dumps or heaps insofar

as they contained data relevant to subaqueous disposai. Similarly, results from marine

disposai operations are only included where such work illustrates principles which are

applicable universally to aquatic systems. The results of this literature review are

summarized below. Section 7.0 provides a comprehensive bibliography of information

on both the theory and practice of subaqueous disposai.

Of all the conclusions drawn from this study, perhaps the most salient is that AMD

poses serious disadvantages for land-based disposai of reactive mine wastes and that the

underwater disposai of such wastes holds considerable promise for suppressing acid

generation. Nevertheless, the potential long-term impacts associated with subaqueous

disposai remain poorly understood.

- Il -

EXECUTIVE SUMMARY

Various factors, sometimes acting synergistically, determine the potential for mine

wastes deposited underwater to generate acid and, consequently, the potential for

biological impacts. These factors include, among others, the natural chemistry of the

receiving environment, physicochemical conditions which may help limit concentrations

of dissolved metals, hydrochemical conditions that may increase heavy metal solubility

and the composition of the mine wastes being deposited. Of the range of predictive

tests available to evaluate potential for acid generation (Section 2.3), the kinetic shake

flask test appears somewhat suitable for subaqueous storage of reactive mine wastes.

The complex processes of bioavailability of metals in lake-bottom sediments and

bioaccumulation in the freshwater food chain are not well understood, particularly with

regard to reactive mine waste disposai. To help improve the level of understanding,

lake studies should be conducted whereby post-depositional reactivity of submerged

wastes is evaluated to determine if benthic effluxes of selected metals, i.e. Cu, Pb, Zn,

Cd, Mn, Fe, As, and Hg are present and to what extent they are obviated by the graduai

deposition of a veneer of natural sediments. Apart from potential impact, other

biological effects of underwater disposai include turbidity, sedimentation on lake

bottoms and toxicity to aquatic organisms.

Following a review of the literature relating to acid mme drainage, subaqueous

disposal, and its potential biological implications, numerous case studies documenting

existing occurrences of subaqueous disposai in a freshwater environment were

reviewed. The cases analyzed within British Columbia include Buttle Lake, Benson

Lake, Babine Lake, Bearskin Lake (proposed), Brucejack Lake (proposed), Kootenay

Lake, Pinchi Lake, Summit Lake, Equity Silver Mines Ltd. (flooded open-pit), Endako

Mine (flooded open-pit), Cinola Gold Project (proposed) and Phoenix Mine. Other

Canadian and U.S. cases examined include Garrow Lake, Northwest Territories;

Mandy Lake, Anderson Lake, and Fox Lake, Manitoba; and Reserve Mining Co. Ltd.,

Silver Bay, Minnesota. Generally it was concluded that although the case studies

reviewed represented a diversity of environments, the results yielded were somewhat

inconclusive. Data were generally superficial, only remotely relevant, reflect

questionable sampling practices, and were not gathered with a view toward better

understanding the long-term impacts associated with the subaqueous disposai of

reactive mine wastes.

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EXECUTIVE SUMMARY

Based on the results of the literature review, it is recommended that field studies be

undertaken to evaluate the post-depositional reactivity of sulphide-bearing mine wastes

and to conduct more detailed, site-specific investigations of potentia1 bio1ogical impacts

in the freshwater receiving environment. Studies should include a detailed evaluation

of the ore and tailings mineralogy, particle size distributions, predicted settling

behaviour of mine wastes, and leaching behaviour or reactivity of the wastes once

exposed to freshwater. It is considered critical that geochemical and limnological field

investigations be completed in concert to both increase our knowledge of the factors

which control metal release or uptake by tailings and the potential associated, direct or

indirect, impacts that might accrue to the biological community.

The geochemical studies recommended require analyzing interstitial waters collected

from suites of cores raised from submerged tailings deposits in a number of lakes

including Buttle Lake, Benson Lake and/or Kootenay Lake, British Columbia; Mandy

Lake, Fox Lake, and/or Anderson Lake, Manitoba. Such studies should embrace a

variety of deposits including unperturbed sediments and tailings with contrasting

mineralogies, and should include assessment of alteration effects and connate water

and/or groundwater chemistry in contrast to tailings disposed on-land. The studies

should include locations no longer receiving mine wastes and active depositional

regimes. It is highly recommended that the geochemical investigations include

chemical analyses of selected major and minor element concentrations in the solid

phases from which the pore waters are extracted, mineralogical characterization, and

measurements of organic carbon concentrations.

Comparative mineralogic studies of bath facies should be undertaken to contrast the

extent and nature of minerai alteration where tailings of the same composition have

been discharged bath underwater and on-land (e.g., Buttle and Benson l..akes). Such

comparisons have the potential to provide a particularly enlightening suite of examples

of the relative diagenetic bchaviour of tailings exposed to the atmosphere versus those

submerged in a freshwater environment.

. .

In association with the geochemical analyses described above, limnological and

biological investigations must be completed to link the complex process of metals

release from submerged wastes to their potential uptake by aquatic organisms and

bioaccumulation in the food chain. The purpose of these studies will be to describe the

lake(s) considered in terms of features that can assist in predicting the impacts of mine

- lv -

EXECUTIVE SUMMARY

wastes deposited in similar lakes. Lake morphology and hydrology, physical and

chemical limnology and biological characteristics should ail be measured to allow

investigators to calculate lake turnover and residence time, determine circulation and

mixing features, and evaluate the potential for wastes to be mobilized and the rate at

which contaminants would be dispersed from the mine waste deposit. A better

understanding of metal transfer between sediments and the aquatic food chain would

also be achieved.

Site-specific experiments on lacustrine biota should be designed to establish the impacts

of heavy metals on hoth infauna and epifauna. Meta! levels within the tissues of these

organisms may reflect metal uptake rates and the potential for bioaccumulation in the

food chain. It is also advised that one or two suitable fish species (i.e. those

characterized by low mobility, long life-spans, and/or higher trophic level feeding) be

chosen for tissue metals analysis due to the high interest in fish by both regulatory

agencies and the general public.

Finally, based on a combination of theory as documented in this literature review, and

empirical field study data, it is recommended that a decision model be developed to

evaluate the suitability of future underwater waste disposai strategies. Initial attempts

at this have been confounded by the insufficient and/ or unreliable data describing

conditions at existing subaqueous disposai sites.

The type of decision mode! proposcd would incorporate physical, chemical,

geochemical, biochemical, limnological and biological conditions, identified in theory

and refined through field investigation, in a critical path framework to evaluate the

environmental implications associated with strategies for the subaqueous disposai of

mine wastes. It would provide a pragmatic method for screening disposai alternatives

by bath industry and government regulators, based on a fatal flaw approach that would

identify key potential problem areas for given proposed discharge strategies. The

decision mode! would be developed based on theoretical and case study information

collected through this review, coupled with empirical data gathered through field

studies such as those outlined above, and would assist both industry in effectively

choosing methods of reactive mine waste disposai and government charged with the

responsibility for ensuring wastes are disposed of in an environmentally acceptable

manner.

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ACKNOWLEDGEMENTS

We would like to acknowledge the British Columbia Acid Mine Drainage Task Force

for their cooperation and support throughout this project. We would specifically like to

acknowledge the following individuals from the Task Force Technical Comittee who

provided valuable comments during the draft review: Mr. J.D. Robertson, P.Eng. -

Chairman; and Mr. K.D. Ferguson, P.Eng. Mr. M. Filion of CANMET and Mr. W.

Fraser of Hudson Bay Mining and Smelting Co. Limited also provided thoughful

insights to the final version of this report.

The Subaqueous Disposai of Reactive Mine Wastes report was completed as a

multidisciplinary exercise through the efforts of those individuals listed below. Their

technical and editorial input have directly resulted in the successful completion of this

document.

• MR. C.A. PELLETIER, B.Sc. (CHEMISTRY)

As Project Manager, Mr. Pelletier oversaw the project and reviewed the final

document. His considerable experience with environmental issues relating to

the mining industry, especially in subaqueous disposai of mine wastes, has been

an invaluable source of information throughout the project. He provided

guidance to the team of scientists accumulating the information for the

compilation of this literature review.

• DR. T.F. PEDERSEN, B.Sc. (GEOLOGY), Ph.D. (MARINE GEOCHEMISTRY)

Dr. Pedersen's extensive experience in the field of sediment geochemistry (he is

Associate Professor, Department o{ Oceanography, UBC), particularly his

research on the chemical reactivity of submerged mine tailing deposits. His

specific research on tailings deposits in Buttle Lake, Rupert Inlet and Alice Arm,

B.C., was of particular importance.

• DR. G.W. POLING, B.Sc. (MINING AND METALLURGICAL ENGINEERING),

Ph.D. (MINERAL PROCESSING)

Dr. Poling is one of Canada's leading experts in the environmental management

of mining operations with special emphasis on tailings disposai, waste solids

-vl -

ACKNOWLEDGEMENTS

handling and acid drainage contrai. He is Professor and Head of Mining and

Mineral Process Engineering, UBC. His environmental work for the Island

Copper mine significantly added to the knowledge base of this work.

• DR. D.V. ELLIS, B.Sc. (ZOOLOGY), M.Sc. (BIOLOGY), Ph.D. (MARINE BIOLOGY)

Dr. Ellis's 30 years of experience in aquatic biology and environmental sciences,

in particular that related to the impact of underwater tailings disposai in the

aquatic environment, provided a significant source of knowledge in the

documentation of limnological and biological components of the study.

Complementing his wealth of practical experience with subaqueous tailings

disposai, Dr. Ellis bas written extensively on the subject, including Mine Tailings

Disposai (Ann Arbor Science Press, 1982).

• DR. D.W. DUNCAN, 8.S.A. (HONS.), Ph.D. (MICROBIOLOGY)

Dr. Duncan has been working in the field of microbiological oxidation of

sulphide minerais since 1962. His work has included research to prevent the

unwanted oxidation of waste which generates acid mine drainage, and bas

included the first practical procedure to estimate the acid generating potential of

sulphide bearing materials. Dr. Duncan's specific involvement in the assessment

of the potential for AMD from mine tailings in Buttle Lake was also of

considerable help. He contributed to the limnological and biological

components of this study.

• DR. T.G. NORTHCOTE, B.A., Ph.D. (FRESHWATER ECOLOGY/LIMNOLOGY)

Dr. Northcote is Professor, Forest Sciences and Zoology at UBC, and specializes

in limnology and the ecology of freshwater habitats. He has worked extensively

in lake ecosystems investigating the relationship of fish ecology and behaviour,

and bas extensive experience with mining-related activities and their associated

effects on the freshwater environment. Dr. Northcote provided bis extensive

knowledge in freshwater biological and limnological systems.

Other members of the Rescan team who contributed to the documentation of

Subaqueous Disposai of Reactive Mine Wastes include:

-vil -

ACKNOWLEDGEMENTS

• Mr. K.P.M. Dushnisky, B.Sc. Hons. (Ecology) M.Sc. (Resource Management)

• Mr. S.T. Robinson, M.A.Sc. (Environmental Engineering)

• Mr. S.R. Bilevicius, M.A.Sc. (Environmental Engineering)

• Mr. R.S. Kinasevich, M.Sc. (Minerai Technology)

Without the dedication and support of the individuals noted above, the successful

completion of this document would not have been possible.

- VIII -

TABLE OF CONTENTS

EXECUTIVE SUMMARY ............................................................................................................. i

ACKNOWLEDGEMENTS .......................................................................................................... vi

TABLE OF CONTENTS ............................................................................................................ .lx

1.0 INTRODUCTION .......................................................................................................... 1-1

1.1 Purpose of Study .............................................................................................. 1-1

1.2 Study Scope and Methodology ...................................................................... 1-2

1.3 Overview of Literature Search ...................................................................... 1-4

1.4 Discussion of Literature on Subaqueous Mine Waste Disposal.. ............ 1-5

2.0 ACID MINE DRAINAGE (AMD) ................................................................................. 2-1

2.1 AMD Generation/Consumption .................................................................. 2-1

2.2 Rates of Acid and Alkalinity Production ..................................................... 2-4

2.3 Acid Mine Drainage - Prediction .................................................................. 2-8

2.3.1 Static - Whole Rock Acid/Base Accounting ................................ 2-8

2.3.1.1 B.C. Research Initial Test ............................................. 2-8

2.3.1.2 Acid/Base Accounting ................................................... 2-9

2.3.1.3 Alkaline Production Potential: Sulphur Ratio ........... 2-9

2.3.1.4 Net Acid Production Test .. ......................................... 2-10

2.3.1.5 Other Static Tests ........................................................ 2-10

2.3.2 Kinetic - Accelerated Weathering Tests .................................... 2-10

2.3.2.1 B.C. Research Confirmation Test ............................. 2-10

2.3.2.2 Humidity Cell Test ...................................................... 2-11

2.3.2.3 Shake Flask Tests ......................................................... 2-11

2.3.2.4 Soxhlet Extraction Tests ............................................. 2-12

2.3.2.5 Columns/Lysimeters ................................................... 2-12

2.4 Summary ........................................................................................................ 2-12

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TABLE OF CONTENTS

3.0 SUBAQUEOUS DEPOSITION .................................................................................. 3-1

3.1 Biogeochemical Zonation - Theory ............................................................. 3-1

3.2 Factors Affecting the Biogeochemical Rubber Band ............................... 3-6

3.3 Diagenetic Consequences .............................................................................. 3-9

3.3.1 Role of Oxyhydroxide Phases ........................................................ 3-9

3.3.2 Authigenic Sulphides ..................................................................... 3-10

3.3.3 Benthic Fluxes ................................................................................ 3-13

3.4 Application to Submerged Sulphide-Bearing Mine Waste Deposits ... 3-15

3.4.1 The Oxidation Problem: Theory ................................................. 3-15

3.4.2 The Oxidation Problem: Practice ................................................ 3-18

3.4.3 Speciation of Metals in Natural Waters ..................................... 3-20

4.0 LIMNOLOGICAL AND BIOLOGICAL CONSIDERATIONS ................................. 4-1

4.1 Microbiology .................................................................................................... 4-1

4.1.1 Sulphide Oxidizers ........................................................................... 4-2

4.1.2 Sulphate Reducers ........................................................................... 4-5

4.2 Macrobiological Effects ................................................................................. 4-5

4.2.1 Turbidity Effects .............................................................................. 4-6

4.2.2 Sedimentation Effects ..................................................................... 4-7

4.2.3 Toxicity Effects ................................................................................. 4-8

4.2.3.1 Phytoplankton ................................................................. 4-8

4.2.3.2 Macrophytes .................................................................. 4-10

4.2.3.3 Zooplankton .................................................................. 4-10

4.2.3.4 Zoobenthos .................................................................... 4-11

4.2.3.5 Fish ................................................................................. 4-11

4.2.4 Contamination Effects .................................................................. 4-14

4.2.5 Receiving Water Conditions ........................................................ 4-15

5.0 SELECTED CASE STUDIES ........................................................................ 5-1

5.1 Introduction ..................................................................................................... 5-1

5.1.1 British Columbia Mines Sites ........................................................ 5-1

5.1.1.1 Babine Lake (Granisle Mines) ..................................... 5-1

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TABLE OF CONTENTS

5.1.1.2 Bearskin Lake, Noramco/ Chevron Minerais Ltd .................................................... 5-5

5.1.1.3 Benson Lake (Coast Copper Co./Cominco) ............. 5-6

5.1.1.4 Brucejack Lake (Newhawk Gold Mines Ltd.) ......... 5-11

5.1.1.5 Buttle Lake (Westmin Resources Ltd.) .................... 5-13

5.1.1.6 Kootenay Lake (Cominco's Bluebell Mine) Riondel (Pb, Zn, Ag) ................................................... 5-21

5.1.1.7 Kootenay Lake (Dragoon Resources Ltd., formerly David Minerais) Ainsworth (Pb, Zn, Ag, Au) ......... 5-24

5.1.1.8 Pinchi Lake (Cominco Ltd.) ....................................... 5-24

5.1.1.9 St. Mary's River/Kootenay River (Sullivan Mine - Cominco Ltd.) ................................. 5-24

5.1.1.10 Summit Lake (Scottie Gold Mines Ltd.) .................. 5-25

5.1.2 Other Canadian Mine Sites .......................................................... 5-27

5.1.2.1 Fox Lake (Farley and Sherridon Mines, Sherritt Gordon Mines Ltd.) ...................................... 5-27

5.1.2.2 Garrow Lake (Polaris Mine, Cominco) .................... 5-31

5.1.2.3 Mandy Lake at Mandy Mine (Hudson Bay Mining and Smelting Ltd., Flin Flon, Manitoba) .................................................... 5-36

5.1.2.4 Anderson Lake (Hudson Bay Mining and Smelting Ltd.) ........................................................ 5-39

5.1.3 International Mines ....................................................................... 5-40

5.1.3.1 Silver Bay, Minnesota, on Lake Superior (Reserve Mining Co.) .................................................. 5-40

5.1.4 Flooded Pits and Shafts ................................................................ 5-41

5.1.4.1 Endako Mines Division (Placer-Dome Inc.) ............ 5-41

5.1.4.2 Equity Silver Mines Ltd. (Placer-Dome Inc.) .......... 5-42

5.1.4.3 Phoenix Copper Mine (formerly with Granby Mining Co.) Decommissioned Open-pit.. ................. 5-44

5.1.4.4 City Resources (Canada) Limited (formerly Consolidated Cinola Mines Ltd.) Queen Charlotte Islands, B.C ..................................... 5-44

5.2 Summary ......................................................................................................... 5-46

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TABLE OF CONTENTS

6.0 CONCLUSIONS AND RECOMMENDATIONS ..................................................... 6-1

6.1 Conclusion ........................................................................................................ 6-1

6.2 Recommendations for Field Investigations ................................................ 6-4

6.2.1 Geochemical Investigations ........................................................... 6-4

6.2.2 Limnological and Biological Investigations ................................. 6-8

7.0 BIBLIOGRAPHY .......................................................................................................... 7-1

APPENDIXA Glossary of Terms

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LIST OF TABLES

Table Page

3-1 Oxidation Reactions of Sedimentary Organic Matter ...................................... 3-2

3-2 Constants for Selected Solubility Equilibria .................................................... 3-16

3-3 Equilibrium Model: Effect of Complex Formation on Distribution of Metals in Aerobic Waters ........................................................ 3-23

5-1 Selected Case Studies, Lake Disposai of Mines Wastes .................................. 5-2

5-2 Selected Case Studies, Flooded Pits and Shafts ................................................ 5-3

5-3 Water Quality Characteristics of Babine Lake in Environs of Granisle Copper Operations ........................................................... 5-4

5-4 Water Quality Characteristics of Babine Lake in the Immediate Environs of the Bell Copper Mine-Milling Operation ................ 5-5

5-5 Estimated Minerais Composition for Coast Copper Mine Tailings .............. 5-8

5-6 Mean Concentrations of Sorne Heavy Metals in Trout

and Water from Benson and Maynard Lakes .................................................... 5-9

5-7 Simulated Metal Concentrations at Outlet of Brucejack Lake After 8 Years of Operating the Tailing Disposai System .............................. 5-14

5-8 Comparison of Tailings Entering Buttle Lake Before and After Commencing Treatment ........................................................................... 5-16

5-9 Dissolved Interstitial Metal Concentrations in Buttle Lake Cores,

Measured by Graphite Furnace AAS (Data from Pedersen 1983) .............. 5-17

5-10 Zn, Cd and Cu Concentrations in Supernatant Water Immediately

After Core Collection (A) and After 8-11 h of Partial Exposure of the Tailings on the Top of the Cores to Air (B) ......................................... 5-18

5-11 210Pb (supported and unsupported) and Zn Concentrationsin Sediments from Core B8 ................................................................................. 5-20

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5-12 Background (A) and Maximum (B) Element Concentration (ppm) and the Ratios of B/A ( =F, a Sediment Enrichment Factor) in a Core from Kootenay Lake Compared with Other Lakes ........................... 5-23

5-13 Deposited Tailings Characteristics in Summit Lake ...................................... 5-26

5-14 Water Quality at Scottie Gold, 1979 ................................................................. 5-27

5-15 Description of On-Land Tailings from Sherridon ........................................... 5-29

5-16 Description of Bore Hole Five (BH5), Sherridon Tailings, Beneath Fox Lake ................................................................................................ 5-30

5-17 Minerai Reactions Associated with the Oxidation of Suiphide Minerais in Reactive Mine Tailings .................................................................. 5-32

5-18 Garrow Lake Surface Water Quality, 1982 ...................................................... 5-36

5-19 Composition of Tailings from Mandy Mine and Cuprus Mine ..................... 5-37

5-20 Soluble Constituents in a 1:5 Distilled Water Leach of Tailings from Mandy and Cuprus Mine ............................................................ 5-37

5-21 Quality of Inlet and Outlet Water at Mandy Lake ......................................... 5-38

5-22 Water Quality in Endako Pit .............................................................................. 5-41

5-23 A.M.D. Treatment Statistics ESML's Water Quality Data(Gallinger, 1988) ................................................................................................... 5-43

5-24 1984 Water Quality Data Phoenix Fiooded/Decommissioned Open Pit ................................................................................................................. 5-45

5-25 Sulphate Ion and Dissolved Copper Concentrations in Providence Lake ................................................................................................... 5-45

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LIST OF FIGURES

Figure Page

2-1 Pourbaix Diagram for the Fe-S-H2O System ..................................................... 2-3

2-2 Pourbaix Diagram for the S-H20 System ........................................................... 2-3

2-3 The Cumulative Acidity for an Acid Producing Coal Sample for Various Leaching Intervals ..................................................... 2-6

2-4 The Daily Acidity Values for an Acid Producing Coal Sample for Various Leaching Intervals ..................................................... 2-6

2-5 The Cumulative Alkalinity for a Limestone Sample for Various Leaching Intervals ............................................................................. 2-7

2-6 The Daily Alkalinity Values for a Limestone Sample for Various Leaching Intervals ............................................................................. 2-7

3-1 Schematic distribution of biogeochemically important species in interstitial waters in sediments showing the zonation typically observed in marine and lacustrine sediments .................................... 3-4

3-2 Schematic distribution with depth of manganese and iron species in interstitial water and solid sediments .............................................. 3-11

3-3 Solubility of selected oxides and hydroxides plotted as free metal ion concentration in equilibrium with solid oxides or hydroxides vs. pH ................................................................................ 3-17

3-4 The concentration of Zn2+ in equilibrium with hydrozincite vs. pH (25 ° C, I = 4 x 10-3, pKSO = 67) .............................................................. 3-19

3-5 Summary of major processes and mechanisms in the interactions between dissolved and solid metals species in surface waters ...................... 3-21

5-1 Benson Lake Showing Contours, Secchi Dise Stations, and Discharge l.ocation ......................................................................................... 5-7

5-2 Benson Lake Thermal Stratification, Station 1 ............................................... 5-10



-XV-

5-5 Garrow Lake - Salinity and Temperature Profiles (Kuit and Gowans, 1982) .................................................................................... 5-34

5-6 Typical Physicochemical Stratification of Garrow Lake by Water Depth (Ouellet and Page, 1988) ..................................................................................... 5-35

-xvi -

1.0 INTRODUCTION

The safe and economic disposai of waste rock and tailings is a continuing source of

difficulty plaguing both the mining industry and governments. When these mine waste

materials contain sulphides, the conventional practice . of on-land containment has

resulted in the generation of highly acidic water and the concomitant leaching of trace

metals. This has far reaching and serious environmental implications and supports

more recent initiatives toward subaqueous disposai, especiaJly since the prohlem may

continue long after mining operations have ceased. To investigatc the problem of acid

mine drainage, several industry-government task forces have been established including

the national Mine Environment Neutra! Drainage (MEND) program and the British

Columbia Acid Mine Drainage Task Force. These committees are playing an

increasingly important role in supporting research, including this literature review on

the Subaqueous Disposa! of Reactive Mine Wastes, to encourage the establishment of

environmentally safe methods for mitigating acid mine drainage.

1.1 Purpose of Study

Subaqueous disposai of mine waste rock and tailings relies on the premise that

submerged mine wastes, minimally exposed to oxygen and bacteriaJ processes, will be

suppressed from generating acid. While this premise may be correct, it remains largely

unproven. A limited number of reviews of the short and long-term envfronmental

implications of the disposai of mine wastes to the aquatic environment have been

undertaken and only a few of the field studies have addressed such issues. Although

basically an unproven method of mitigation, there is an ever-increasing number of

mines being developed which are proposing to adopt subaqueous disposai as a means of

combating potential acid generation problems. A thorough review of the history of

existing subaqueous disposai practices is, therefore, needed in order to provide a

foundation for making informed and confident judgements on the suitability of aquatic

discharge strategies in the future.

Responding to the need for economic and effective approaches to resolve the problem

of acid mine drainage, the British Columbia Acid Mine Drainage Task Force issued a

call for proposais to review the British Columbia experience with subaqueous disposa}

1 - 1

INTRODUCTION

of reactive mine wastes in a freshwater environment. Briefly, the terms of reference for

the study were as follows:

• The consultant was to investigate the effectiveness of disposing of waste rock

and tailings underwater to discourage acid generation, and to assess the

potential impacts of this disposai method on the freshwater environment.

• Based on a review of relevant literature, including project specific data gathered

from British Columbia mines utilizing subaqueous disposai, the consultant was

to evaluate the biological impacts of mine waste disposai in a freshwater

environment, considering both the physical and chemical effects of subaqueous

deposition.

Following a review of the proposais submitted, the British Columbia Acid Mine

Drainage Task Force commissioned Rescan Environmental Services Ltd. in

November 1988 to complete the review contained herein.

1.2 Study Scope and Methodology

This study was completed based on a comprehensive literature review of ail aspects of

subaqueous disposai of reactive mine wastes in a freshwater environment. Though

emphasis was placed on the British Columbia experience as mandated in the RFP,

consideration was also given to mines utilizing subaqueous disposai practices in other

parts of Canada and the United States. Numerous sources of information were

consulted during the review through bath desk and computer-aided search methods.

The types of literature investigated included, among others, information on acid mine

drainage mechanisms, aquatic chemistry, geochemistry, hydrogeochemistry,

biochemistry, microbiology, limnology, and aquatic ecology. While the scope of this

review is limited to freshwater environments, land-based tailings, leach dumps/heaps,

and submarine disposai systems were also considered insofar as they provided data

relevant to freshwater disposai.

This review, although contemporary and considered to be timely, is certainly not the

first to address the disposai of mine wastes. Other reviews, studies and reports, some

with quite comprehensive literature reviews, have appeared over the past two decades.

In fact, a 1970 report entitled The Disposai of Mining and Milling Wastes with Particular

1 - 2

INTRODUCTION

Reference to Underwater Disposai was somewhat similar in scope to the one submitted

here. The fundarnental difference between past efforts and this attempt to address

subaqueous disposal is that this report evaluates the state-of-the-art in underwater

disposai theory, coupled with data gathered from mines utilizing the practice, bath to

advance the current level of understanding and recommend practical strategies for

future field studies to improve our knowledge of the long-term behaviour and impacts

of extant and abandoned subrnerged tailings deposits.

This report is cornprised of two parts: a theoretical literature review and an overview of

selected case studies. The theory component first describes fundamentals of acid mine

drainage that have plagued land-based tailings disposai, then examines the practice of

subaqueous deposition with emphasis on physical, chemical, geochemical,

microbiological and macrobiological implications for the freshwater environment. The

literature review supporting this theoretical section is extensive and provides a

comprehensive database for those interested in reactive mine waste disposai.

The case studies selected were chosen based on a number of criteria including data

accessibility and method of subaqueous disposai (flooded tailings ponds, pits,

underground workings). Along with the British Columbia experience with freshwater

subaqueous disposai, Rescan also reviewed mines in other parts of Canada and the U.S.

The sites examined within British Columbia include Buttle Lake, Babine Lake, Benson

Lake, Brucejack Lake (proposed), Bearskin Lake (proposed), Kootenay Lake, Pinchi

Lake, Summit Lake, Equity Silver Mines Ltd. (flooded open-pit), Endako Mine

(flooded open-pit), City Resources (Canada) Limited - Queen Charlotte Islands and

Phoenix Mine. Outside British Columbia, the sites investigated include Garrow Lake,

Northwest Territories; Mandy Lake, Anderson Lake and Fox Lake, Manitoba. Also

reviewed was Reserve Mining Co. Ltd., Silver Bay on Lake Superior, Minnesota.

Based on a review of theoretical considerations and the results of the case study

analysis, conclusions are drawn and recommendations are made for future field

investigations to increase our understanding of the long-term behaviour and impacts of

subaqueous mine waste disposai through the collection and analysis of empirical data.

This improved understanding will permit more confident review and evaluation of

future subaqueous disposai strategies.

1 - 3

1.3 Overview of Literature Search

INTRODUCTION

A comprehensive search was made for literature dealing with the subaqueous disposal

of reactive mine wastes, and included consideration of hydrogeochemistry,

microbiology, biochemistry, biology and limnology. The bibliography compiled in

Section 7.0 represent a substantial collection of scientific and technical knowledge on

underwater tailings disposal.

Although the scope of this report focusses on subaqueous tailings disposai, the

literature search located and incorporated documentation on land tailings disposal and

leach <lumps, if these were found to contain useful information.

The literature review incorporated on-line computer database searches using a variety

of index systems, in addition to an extensive effort to locate information in journals and

technical textbooks. Lists of keywords and authors were compiled from an overview of

key publications. Database systems used include the comprehensive CAN/OLE and

DIALOG systems, and the QL System. Recognized scientific and engineering index

systems that were searched include:

• NTIS (1964 - present) - listing U.S. Government Publications, although not a

complete listing of U .S. Environmental Protection Agency documents;

• Pollution Abstracts (1970 - present) - a leading source of literature on

environmental quality, solid wastes and water pollution;

• Chem Abstracts (1967 - present) - citations of ail research into chemistry and its

applications;

• Engineering Index - COMPENDEX (1970 - present) - abstracted information

from world-wide engineering and technical literature including over

4,500 journals, government reports and books;

• Enviroline (1971 - present) - covers environmental information and indexes over

5,000 international publications, including geology, biology and chemistry;

Aquaref-WATDOC (1970 - present) - Canadian water resources references

encompassing scientific and technical literature;

1 - 4

INTRODUCTION

• Mintec/Minproc - CANMET (1968 - present) - Canadian Mining and Mineral

Processing index.

Subsequent to computer database searches, publications were retrieved from the

following collections:

• University of British Columbia Library;

• Geological Survey of Canada Library;

• Environment Protection Service (Environment Canada) Library, Pacifie Region.

• Fisheries and Oceans Library;

• Inland Waters Directorate (Environment Canada);

• Vancouver Public Library, Sciences and Technology Division;

• University of Waterloo Groundwater Research Library;

• Rescan Environmental Services Library, and the personal libraries of associates

who have contributed to the study.

1.4 Discussion of Literature on Subaqueous Mine Waste Disposai

Much bas been observed, investigated and written on acid mine drainage and its

control, acid generation within terrestrial tailings impoundments and waste dumps, and

oxidation-reduction reactions involving discarded iron and other residual metallic

sulphide minerals (pyrite, pyrrhotite, arsenopyrite, chalcopyrite, and residual zinc and

lead sulphides). Only a small portion of the literature attempts to address underwater

disposal in the freshwater environment. Few reports contain data or observations from

existing operations which are sufficient for general evaluation of freshwater disposal.

Pertinent references strongly suggest that optimum conditions for subaqueous disposal

are not fully understood and that more analyses will be required to permit developing

guidelines for its use.

1 - 5

INTRODUCTION

Sorne of the more pertinent references include:

• Bohn et al., 1981 • Kuit, 1982• Daley et a1., 1981 • Kuit and Gowan, 1982• Halbert et al., 1982 • Littlepage et al., 1984• Hamilton and Fraser, 1978 • McKee et al., 1982. Hawley, 1975 • Pedersen and Losher, 1988• IMPC, 1982 • Pedersen, 1983• Kennedy and Hawthorne, 1987 • Ripley et al., 1978• Knapp, 1981 • Salomons and Forstner, 1988

In addition, the content or themes of several biologically oriented overviews or articles

suggest caution about underwater tailings disposai. Sorne references of this type that

require special attention are:

. Allan, 1986 • Nriagu, 1979, 1989• Campbell et al., 1988 . Pharo, 1979• Daley et al., 1981 • Reynoldson, 1987• Evans and Lasenby, 1988 • Roch et al., 1982, 1985• Forstner and Wittmann, 1983 • Schindler, 1987• Hamilton, 1976 • Sexton, 1963• Harvey, 1976 • Sly 1977, 1984• Jackson 1978, 1980, 1988 • Van Duyn - Henderson• Malo, 1977 and Lasenby, 1986

There is a substantial body of literature on marine disposai, but judicious consideration

is required when this information is applied to freshwater subaqueous disposai. Little

information bas been located on flooded pits and mine shafts (Gallinger, 1988). ln

British Columbia there are several depleted and closed-down workings which now exist

as abandoned or decommissioned open pits.

1 - 6

2.0 ACID MINE DRAINAGE

Acid mme drainage (AMD) is the hydrogeologic expression of a hast of complex

oxidation reactions of sulphide minerals. In general, whenever sulphide minerals such

as pyrite, FeS2, marcasite, FeS2, or pyrrhotite, Fen_1Sn, are exposed to water and to

oxygen, the production of hydrous ferrous sulphate will result. Ground waters

percolating through or flushing around or saturating these minerais will dissolve the

ferrous salts and acid. The ferrous iron will subsequently be oxidized to ferric ions or

ferric oxyhydroxides either chemically or biochemically. A buildup of ferric ion will

usually mean that this species will far exceed oxygen as the dominant oxidizing agent of

exposed sulphide minerals. As the reactions progress, the rate of oxidation of the

sulphides continues to increase as more ferric ion is generated and this in turn oxidizes

more iron sulphide to release more ferrous and eventually more ferric ions. In this

sense the oxidation of the sulphides can become auto-catalytic.

Percolation or flushing of surrounding ground waters, which picks up the reaction

products of sulphide oxidation, will normally contain some alkalinity, such as CaC03,

which can react to neutralize the acid. Thus the eventual drainage might be neutral in

pH with elevated sulphate concentrations. Whether or not the percolating or flushing

waters will eventually emerge as "acid drainage", therefore, will depend not only on the

balances of acid-producing sulphides and acid-consuming rocks (such as CaC03) but

also on the rates of acid production versus alkaline production.

Drainage conditions can change dramatically over long periods of time. Should

carbonate be consumed, production of neutralizing alkalinity would cease while acid

production continued. A neutral drainage could then rapidly convert to a strongly acid

one. It should not be surprising, therefore, that predictive technology of AMD is

necessarily complex.

2.1 AMD Generation/Consumptlon

Thermodynamically, sulphides are only stable in reducing aqueous environments. One

of the best ways to depict the various domains of thermodynamic stability is to use an

Eh versus pH diagram. Cloke (1966) concisely outlines methods for construction of

Eh-pH diagrams. Figure 2-1 shows such a "Pourbaix Diagram" for the Fe-S-H20 system

2 - 1

ACID MINE DRAINAGE

under standard conditions of temperature (298 ° C) and pressure (1 atmosphere) for 10--0

activity of dissolved iron species (Garre)s and Christ, 1965).

Figure 2-1 indicates that oxidizing conditions (higher Eh) are required to dissolve pyrite

to yield Fe++ or eventually Fe+++ species. A Pourbaix diagram for the S-H20 system is

shown in Figure 2-2 which also indicates the need for oxidizing conditions to form S04

=

ions.

Commonly accepted reactions describing the natural oxidation of pyrite are as follows

(Singer and Stumm, 1969):

FeS2

+ 7 /2 02

+ H20 --> Fe++ + 2so

4-- + 2H+

Fe++ + 1/4 02

+ H+ --> Fe+++ + 1/2 H20

FeS2

+ 14 Fe +t + + 8 H20--> 15 Fe++ + 2s0

4-- + 16H+

Fe+ ++ + 3 H20 --> Fe(OH)

3 (solid) + 3H+

(1)

(2)

(3)

(4)

Singer and Stumm (1969) reported that oxidation of Fe++ (reaction (2) above) was rate

limiting and very slow under sterile conditions. Bacteria, which are ubiquitûus in these

systems, are capable of catalyzing this iron oxidation and hence propagate the leaching

cycles. Oxygen remains a critical reactant by serving to regenerate the ferric ion.

Oxygen is also critical to the growth of bacteria such as Thiobacillus fe"ooxidans and

Thiobacillus thiooxidans which participate in the sulphide oxidation. These

microorganisms normally depend also on dissolved C02

as their source of carbon and

require nitrogen and phosphorus for chemosynthesis and growth. Hence aeration of

water is crucial to the formation of AMD by oxidation of sulphide minerais.

Although some authors attempt to describe acid formation as either a chemical or

electrochemical reaction mechanism, Figure 2-1 indicates that almost every reaction in

the Fe-S-H20 system involves electron transfer or change in oxidation state or

electrochemistry. On the Eh vs. pH diagrams, only vertical lines separating domains or

"fields of stability" involve no electron transfer. Since essentially all lines, except the

equilibrium between Fe3+ and Fe20

3 are either horizontal or sloping, all reactions

depicted in Figure 2-1 are electrochemical in nature.

The prospect of driving the oxidation of sulphides by galvanic coupling between

contacting minerai phases of differing potentials is unlikely to be significant except in

2-2

Eh

+- 1.0•

t0.8

+0.6

+0.4

+0.2

0.0

-0.2

-0.4

-0.6

-0.8

-1.0

2 4 6 8 10 12 14

pH

Figure 2-1 Pourbalx dlagram f()( the Fe-S-H20 systemat 200·K and t atm total pressure. Total dlssolved sulfur specles - 10·1 M. (AfterGarrels and Christ, t 965.)

Eh

+ 1.0·,

+ 0.8'·

+0.6 .............

.......

+0.4

+0.2

0.0

-0.2

-0.4

-0.6 .... ....

'

-0.B

-1.0

2 4 6 8 10 12 14

pH

Figure 2-2 Pourbalx dlagram for the S-H20 system at298 • K and 1 atm total pressure. T otaJ dlssolved sulfur specles = 10·1 M. (AhevGarrels and Christ, 1965.)

ACID MINE DRAINAGE

relatively massive sulphide deposits. In any case transportation of oxidant rather than electron-transfer reactions generally contrais the rate of acid generation as explained further in Section 2.2.

Acid-consuming reactions can involve direct reaction of acid with carbonate minerals

such as calcite or dolomite, i.e.:

or reaction of acid with dissolved carbonate-bicarbonate species in the percolation or flushing of surrounding waters. Factors affecting carbonate solubility include pH, partial pressure of CO2, the surface area of the exposed carbonate mineral and the

nature of the carbonate mineral. Equation (5) predicts that 3.12 grams of calcite would

neutralize the acid generated from the oxidation of 1 gram of pyrite. By analogy 2.87 grams of dolomite (Ca Mg(CO

3)2

) would neutralize the acid from 1 gram of pyrite.

Note that according to equation (5), two moles of dissolved SO/ might persist indicating that acid bas been produced by the oxidation of pyrite. If calcite is present in

the rocks, the SO4 = concentration might be limited to -400 ppm since above this concentration gypsum would be precipitated.

Thcrc is evidence that some carbonate minerais, such as siderite, FeCO3, would have

little or no neutralizing action on acidic waters:

2H+ + FeCO/s)-> Fe++ + CO/g) + H2O (6)

Subsequent oxidation of Fe++ to Fe+++ and precipitation of Fe(OH)is) again releases

H+ ions (see equations (2) and (4)); hence no net H+ ions would be consumed. Thus, distinguishing the nature of the carbonate minerai present can be important.

Weathering of silicates also consumes acid but such reactions are generally too slow to be considered effective alkalinity producers for the prevention of AMD. Olivine is one

exception; neutralization of concentrated acids can take as little as two hours using crushed olivine (Pietersen et al., 1988).

2.2 Rates of Acid and Alkalinlty Production

Although the chemistry, biochemistry and electrochemistry of acid generation processes

are certainly important, the physics of acid/alkalinity generation processes are usually

2-4

ACID MINE DRAINAGE

pre-eminent. The kinetics of acid formation can sometimes be limited by the rate of transportation of oxidant (either 0

2 or Fe+++) to the sulphide surface. There is also

strong evidence that for above-ground mine wastes, the concentration of acidity is a function of the length of time between periodic flushing of the rock (Caruccio and Geidel, 1984). Figures 2-3 and 2-4 show results of simulated weathering studies on pyrite-containing coal and shale. Figure 2-3 shows that increasing the interval between flushing did not produce a decrease in cumulative acidity; instead the cumulative acidity remained relatively constant. Figure 2-4 shows that reducing the frequency of flushings simply increased the concenlralion of acidity in each flush. These results indicate that oxidation of the pyrite continued unabated during periods of relative dryness or flushing.

On the other band, with contained limestone, the higher the frequency of flushing the greater was the cumulative alkalinity produced (see Figure 2-5). Figure 2-6 shows that the concentration of alkalinity produced in these simulated weathering tests was independent of flushing frequency. This suggests strongly that alkalinity produced by dissolution of CaC0

3 in water rapidly attained an equilibrium governed by solubility

considerations. The kinetics of acid and alkalinity production in waste rock above ground and subject to alternate drying and flushing sequences can thus differ dramatically and be highly time dependent.

In contrast to the ahove, storage of reaclive wastes under water produces dramatically different reaction conditions. With atmospheric oxygen as the ultimate crucial oxidant in the acid-producing reactions, underwater disposai can drastically curtail reaction rates. Firstly, the concentration of dissolved oxygen in water can attain a maximum of 8.6 x 10-6 g/cm3 at 1 atmosphere air pressure; far lower than the oxygen concentration in air. Secondly, the diffusion constant for oxygen in water (-2 x 10-6 cm2/sec) is nearly five orders of magnitude less than the diffusion constant for oxygen in air (0.178). Thus less than a metre of relatively stagnant water will reduce the oxidation rate of submerged pyrite effectively to zero. Natural levels of alkalinity in the water body can serve to neutralize a small amount of emergent acidity.

Current theory and limited field data thus support the proposition that storage of fresh but reactive sulphide waste rock or tailings underwater should prevent the generation of acid. Oxygen serves as a critical reactant for the electrochemical oxidation of s-2 to s+ 6

(in S0/) and for maintaining an aerobic bacterial colony to catalyze oxidation of Fe++

2-5

--,,,

0

1 Â 0

400 •

'.,,1

• Cl •• E

& •

>-•

1-- 300 & •

• C

ë ü •

• lu

•••>

1-- 200 •

<l •• ...J • & :::::, • :E • C

• Doily:::::, u •

100 ... • Weekly

• c Bi-monrhly

•

�1 1 1 1 1

10 20 30 40 50 60

DAYS

Figure 2-3 The cumulative acidity for an acid producing coal sample for various leachlng intervals. After Carucclo, F.T. and Geidel, G. (1984)

--150

l Il) C

E C Cl

'.,,1 C

Ot 100 E-

50 j • • •Daily

>-1-- ... • Weeklyë ... À ... u ... c Bi- monthly <l

• • • • • • • • • • • • • •

... • • • • • • •• • • • C •• •• •• •

10 20 30 40 50 60 OAYS

Figure ?-4 The daily acidlty values for an acld produclng coal sample for varlous leachlng lntervals.Atter Carucclo, F.T. and Geidel, G. (1984)

(E§)

>�

z

...J <l: � __J <l:

w > i= <l: __J ::, � ::, u

>-�-z

...J c:x: � ...J

<l:

100 -

80 -

60 -

-"' E Cl 0

40 -0

' "' OI E

20 7

• • • •

• •

• •

• • •

•

• ... ... . ... ...

. ... ... . ... . ... ...

• ... ... Cl Cl

• ... ti Cl

20 40

••

•• • •

••••

• •

... ... ...... ... ..

Cl Cl

60

•• •

... ...

Cl

••

......

80

......

Cl

100

• Daily

... Weekly

Cl Bi-monthly

120

Figure 2-5 The cumulative alkallnity for a limestone sample for various leaching intervals.

---"'E OI 0

0

' "' Cl E

Alter Caruccio, F.T. and Geidel, G. (1984)

15 � A

10

5-

l--1•..:i .... "' ... �.�··· .otl9 ... •1ib.e411.�·�· J.• \. .....1 i 1 1

20 40 60 80

• Daily

... Weekly

Cl Bi-monthly

!iil

100

Figure 2-6 The dally alkal lnity values for a. llmestona sample for varlous leachlng lntervals. After Carucclo, F.T. and Geldel, G. (1984)

CE.§

ACID MINE DRAINAGE

to Fe+++ . This does not, however, suggest that all sulphide wastes can be safely stored

underwater. Should the sulphides be heavily pre-oxidized, then their oxidation products

might be rapidly hydrolyzed to form acids if disposed of underwater. Once the soluble

oxidation products have been dissolved (and hopefully neutralized) oxidative leaching

of residual sulphide should be curtailed as would fresh sulphides.

2.3 Acid Mine Drainage - Predlctlon

Predicting the quality of ground water draining from a mine waste dump or derived

from the supernatant water above a mine waste sediment depends critically on an

assessment of the potential of the waste to produce acidity versus the potential to

produce alkalinity. If a macro environment is closed or isolated, analyses of total acid

production potential versus total alkalinity potential of a waste material should indicate

whether the system will remain neutral or become acidic over a geologically long period

of time. So called static methods of acid/base accounting might then have merit at

least as screening tests to determine whether more expensive kinetic tests are

warranted.

ln most disposai systems water and dissolved gases will percolate or flow in and out of

the micro-environments generating acidity and/or alkalinity. The rates at which

sulphides produce acidity and calcareous components produce alkalinity are not equal.

Thus the kinetics of acid/ alkaline release become important. Several predictive kinetic

tests have been developed ta attempt ta accelerate weathering reactions and enable

direct determination of aqueous effluent or supernatant qualities.

2.3.1 Static - Whole Rock Acid/Base Accounting

Simplicity, rapidity and low cost makes static-whole rock determinations of acid

production potential versus alkaline production potential popular predictive methods.

There are several published procedures for determining these characteristics of waste

mate rials:

2.3.1.1 B. C. Research Initial Test

This test uses direct sulphuric acid titration of -400 mesh ground material ta an end

point of pH = 3.5 and room temperature ta determine the acid consumption for

neutralization potential. Acid production potential is calculated by analyzing the

2-8

ACID MINE DRAINAGE

sample for total sulphur content and assuming that all of this would eventually be

converted to s+6 in H2SO

4• Both acid production potential (AP) and acid consumption

or neutralization potential (NP) are expressed as equivalents of kg CaCO3

per tonne of

sample. If AP > NP then the waste rock or tailing is said possibly to generate AMD

(Duncan and Bruynesteyn, 1979).

In some samples, use of total sulphur rather than only sulphide sulphur analyses can

dramatically overestimate acid production potential. Thus the estimate can be very

conservative.

2.3.1.2 Acid/Base Accounting

This test is somewhat similar to the above procedure but uses a hot excess HCl reaction

followed by NaOH titration of excess HCl to an end point of pH 7.0 to determine

"neutralization potential" (NP). AP is determined as above by total sulphur analyses. If

NP-AP is less than -5 kg CaCOitonne then the waste is considered potentially acid

producing (Sobek et al., 1978).

This procedure can overestimate AP for the same reasons as above. The hot acid

titration can also overestimate NP by virtue of measuring and equating non-neutralizing

carbonates such as siderite with effective alkaline producers such as limestone or

dolomite.

2.3.1.3 Alkaline Production Potential: Sulphur Ratio

Alkaline production potential (APP) is determined by measuring the HCl consumed in

two hours of room temperature reaction (with -23 µm sample) after titrating excess acid

added with NaOH to an end point of pH 5.0. APP is calculated as mgs CaCO3

per

500 g sample. AP is determined again by total sulphur analyses and expressed as a

percent. Wastes of low APP:S ratio are suspected of being producers of AMD

(Caruccio et al., 1981).

Comparisons of the APP:S ratio with kinetic weathering charts are claimed to qualify

these ratios as having a predictive capacity. Without such comparisons, the results are

suspect.

2-9

2.3.1.4 Net Acid Production Test

ACID MINE DRAINAGE

This method, developed at Coastech Research, uses hydrogen peroxide addition to

oxidize sulphides in one hour. Acids generated by this oxidation are either consumed

by alkaline constituents or excess acid is titrated to pH 7 using NaOH. Excess Net AP

is then expressed as kg CaCO3

equivalent per tonne. A positive Net AP is believed to

indicate the potential for AMD (Albright, 1987; Lutwick, 1987).

The test accuracy for tailings and waste rock bas been confirmed by field comparison.

2.3.1.5 Other Static Tests

A so-called Hydrogen Peroxide Test (Finkelman and Giffin, 1986) uses hydrogen

peroxide oxidation to determine the amount of pyrite or reactive sulphide in a sample.

The rate of change of pH is compared to a standard curve developed using

pyrite-seeded standards.

Preliminary results appear to be too inaccurate to be of real use.

A Manometric Carbonate Pressure Analysis (Evanglon et al., 1985) is used to

characterize the nature of alkaline-producing carbonate in a sample. Carbonate

contents can be characterized as calcite, dolomite or siderite to assist in evaluating the

effectiveness of alkalinity production in a sample.

Results to date are inconclusive in establishing the reliabi1ity of this technique.

2.3.2 Kinetic - Accelerated Weathering Tests

Leaching and simulated weathering tests are used to develop kinetic data aimed at

characterizing drainage from waste rock or tailing disposai schemes. There are at least

six different published kinetic tests.

2.3.2.1 B.C. Research Confirmation Test

This is a biologically-inoculated oxidation test designed to determine whether sulphide

oxidizing bacteria can generate more acid than can be consumed by an equal quantity

of the sample. The pulp is first acidified to pH 2.0-2.4 and then inoculated with

Tlziobacillus fe"ooxidans culture adapted to grow on pyritic ore. Oxidation initially

2 -10

ACID MINE DRAINAGE

proceeds until a stable pH is reached, then an equivalent weight to the original sample

is added in two increments after 24 and 48 hours. If the pH is less than 3.5, 24 hours

after each addition, the sample is considered to be an acid producer (Duncan and

Bruynesteyn, 1979).

This test has correctly predicted many field results. The initial acidification of the

sample is believed, however, to make some results unrealistic.

2.3.2.2 Humidity Cel/ Test

The humidity cell models the processes of geochemical weathering wherein a bed of

crushed rock or tailing sample is subjected to three days of dry air, then three days of

moist air then one day of leaching in water. Leachates are analyzed for a range of

parameters such as pH, redox, acidity, alkalinity, sulphate, conductivity and dissolved

metals. This test generally takes 8-10 weeks to complete (Sobek et al., 1978).

The humidity cell has correctly predicted field results on tailings and on waste rock

samples. The test seems well suited to waste rock dumps alternately subjected to

infiltration, drying and flushing sequences. Improvements to the original Sobek type

humidity cell are currently being evaluated.

2.3.2.3 Slzake Flask Tests

Tailings or waste rock are ground to -50 µm, washed with sulphuric acid to remove

residual alkalinity, and then subjected to leaching in an inoculated, incubated shake

flask test extending up to three months. The objective is to determine the rate of pyritic

sulphur oxidation by measuring sulphate production in the leachate versus time. In the

original procedure employed by Halbert et al. (1983) a triple factor x two level factorial

design procedure was used to determine the effect of temperature, initial pH and pyrite

oxidizing bacteria on sulphate generation. In addition to pH and sulphate analyses,

dissolved metals were also determined on the leachates produced. Low initial pH

(pH = 3.0) and higher incubation temperatures (21 ° C) resulted in accelerated sulphate

generation rates.

The shake flask tests have correctly predicted field behaviour of tailing samples, but

have not been entirely successful in predicting field behaviour of waste rock.

2 - 11

ACID MINE DRAINAGE

Based on the constant leaching action of the shake tlask test it should not perhaps be

surprising that this test seems accurate in predicting the field behaviour of tailings,

which are often saturated or of low oxygen permeability. The test does not appear to

perform well for waste rock dumps which would be subjected to alternate infiltration,

wetting and drying followed by flushing events.

2.3.2.4 Soxhlet Extraction Tests

A special Soxhlet extraction apparatus is used to simulate an accelerated geochemical

weathering using either acetic acid or distilled water as extractant. The

Singleton-Lavkulich procedure uses an extraction temperature of 68 • C and both acetic

acid and distilled water, while the modified Sullivan-Sobek procedure uses 27 • C and

only distilled water (Singleton and Lavkulich, 1978; Sullivan and Sobek, 1982).

Soxhlet extractions using water have correctly identified acid generating behaviours in

tailings samples and in waste rock samples. The acetic acid extraction procedure

appears to be unrealistic in evaluating AMD behaviours.

2.3.2.5 Columns/Lysimeters

Column/lysimeter studies have been conducted by CANMET (Ritcey and Silver, 1982).

In this testwork, samples were placed in columns or boxes and periodically leached with

water. A simulated rain cycle plus timed intervals of light and darkness were used to

accelerate weathering times � nine-fold. Such tests sometimes last two to three months.

Preliminary results from the CANMET lysimeter tests appear to parallel field

experience closely. Fifty kilogram samples were used, and the tests simulated eight

years of natural Ieaching.

2.4 Summary

ln the absence of oxygen, the bacterially-mediated acid-generating oxidation of sulphide

minerais is strongly inhibited. Under conditions of low turbulence, a relatively thin

layer of water appears to be sufficient to lower the oxidation rate of submerged pyrite

substantially. Diagenetic factors which can promote or inhibit oxidation in sediments

are discussed in the following chapter.

2 -12

ACID MINE DRAINAGE

Although many different methocls exist to predict acid mine drainage, only a few seem

to be accurate for a broad range of tailings and waste rock from metalliferous mines in

Canada. Due to their low cost, simplicity and rapid results, static, whole rock acid/base

accounting procedures witl probably persist for preliminary screening purposes. The

available evidence indicates that acid production potentials should be based on

sulphide sulphur analyses rather than total sulphur. There is also evidence to suggest

that cold acid titrations are preferable, as they exclude carbonates such as siderite

(FeCo).

Certain kinetic tests, although time consuming and expensive, do appear capable of

accurately predicting AMD. Alternating wetting, drying and flushing cycles, such as

humidity cells or lysimeters appear to be the preferable method for evaluation of AMD

potential in unsaturated waste rock stored on land. Shake flask tests appear more

suited for evaluating the acid-generating potential of tailings or waste rock stored under

water.

2-13

3.0 SUBAQUEOUSDEPOSITION

Natural lacustrine sediments characteristically exhibit a chemical zonation with depth

which reflects the integrated influence of physical, microbiological and inorganic

chemical phenomena. Factors as diverse as the reactivity and rate of accumulation of

organic matter, the concentration of orygen in bottom water, input to the sediments of

detrital oxyhydroxide phases, and the presence or absence of benthic fauna all play

major roles in governing the distribution of a number of dissolved constituents in

interstitial water and the distribution of specific authigenic phases at various sub-

bottom depths. AII these factors will, to varying extents, also affect the diagenetic

behaviour of submerged mine tailings. In the following section, therefore, such

diagenetic elements will be described individually as they pertain to natural sediments,

and their collective theoretical influence on the reactivity of submerged

sulphide-bearing materials will be discussed.

3.1 BiogeochemicalZonation-Theory

Bacterial oxidation of reactive organic matter in lacustrine sediments proceeds via a

series of overlapping, enzyme-mediated electron-transfer reactions in which the

thermodynamically-unstable reduced carbon compounds serve as electron donors and

various oxidants act as terminal electron acceptors as degradation proceeds.

Heterotrophs act simply as catalysts during such degradation; they are unable to carry

out reactions which are not thermodynamically possible (Fenchel and Blackburn, 1979).

During oxidation, organic matter will donate electrons to orbitals of lowest available

energy level, as this produces the greatest free energy gain per unit of organic material

oxidized. The reaction sequence (Table 3-1) thus proceeds in an order which is

determined by net free energy yield, with aerobic oxidation, the highest-yield reaction,

preceding (in thermodynamic order) denitrification, manganese and iron oxyhydroxide

reduction, sulphate reduction, and methanogenesis (CO, reduction) (Froelich et

a1.,7979).

In sedimentary systems which are receiving a constant input of reactive (i.e. degradable)

organic detritus, a steady-state zonation will be established with depth if the bacterial

demand for oxidants exceeds the rate of supply by diffusion or advection. In such cases,

3-1

SUBAQUEOUS DEPOSITION

Llsted ln order ol the lree energy ylelds shown, complled lrom Froellch et al. (1979), Bender end Heggle (1984) and

Kadko et el. (1987). The free energy yletds are preeented ar kJ per molc of CX.! o!C!1cd, and the organlc matler

stoichiometrychosen ls equlvalentlothe Redlleld ratlo lor martne planlfon (Redfleìd' 1958)"

Table 3-l

Oxidation Reactions of Sed¡mentary Organic Matter

2

3.

4"

1. Aerobic oxidation: AG " = 475 kJ/mol

(CH2O)106(NH3)16(H3PO4) + t3e O¿ = 1OO CO, + 16 HNO3 + HrPOo + 122H2O

Nitrate reduction (denitrification): AG' = 448 kJ/mol

(CH2O)106(NH3)i6H3POo) + e+.8HNO. = 106CO, + 42'4Nz+ 16NHs + H3PO4

+ 148.4 HrO

Manganese oxide reduction: A G' = -349 kJ/mol

(CH2O)I'6(NH3)16(H3PO.) + eso MnO, + 472H+ = 236 Mn2* + 106 CO2 + I N,

+ H.PO4 + 366 HrO

lron oxyhydroxide reduction: A G' = -114 kJ/mol

(CH2O)106(NH')16(H3PO o) + tz+FeOOH + 848 H+ = 424Fe2' + 106 CO2 + 16 NHg

+ H.PO. + 742H"O

5 Sulphate reduction: A G' = -77 kJ/mol

(CH2O)106(NH3)16(H3PO.) + so SO42' = 106 CO2 + 16 NH. + 5g 52' + H.PO. + 106 HrO

6 Methanogenesis (fermentation): Â G' = '70 kJ/mol

(CH2O)lo6(NHs)16(HsPOa) = 53 CO, + 53 CH4 + 16 NHs + H.PO.

3-2

SUBAQUEOUS DEPOS¡T¡ON

02 will be depleted at some depth below the sediment-water interface. As the oxygen

concentration decreases to low (but not yet zero) levels, nitrate reduction will

commence. The other oxidants listed in Table 3-1 will subsequently be reduced in the

order shown should there continue to be a deficiency of oxidant supply relative to

demand.

In sediments in mesotrophic or eutrophic lakes, the organic carbon content is usually

sufficient to establish anoxic conditions at depths ranging from a few millimetres to one

decimetre. The resulting chemical zonation is therefore characterized by a steadily

decreasing redox potential with depth, and the release to interstitial solution (pore

water) of a number of reaction products (Table 3-1, Figure 3-1). Because the depth to

the oxic-anoxic boundary is determined by relative rates of supply and consumption of

oxidants, the zonation shown in Figure 3-1 can be thought of as a "biogeochemical

rubber band" which is stretched when oxidant demand decreases relative to supply, and

compressed when the opposite conditions occur. Thus, in the absence of secondary

dissolution/ precipitation reactions, and because the oxidant reaction sequence is based

on thermodynamics, the relative distribution of dissolved species in pore water is

constant from site to site; only the clepth scale changes based on the intensity of

diagenesis (Bender and Heggie, 1984). Note that the zones defined in Figure 3-1

overlap in all cases to some extent. In theory, thermodynamic considerations prohibit

such overlaps but they occur for two main reasons in the natural environment. First,

kinetic effects (such as rates of abiologic oxidation of diffusing reduced species) tend to

smear the thermodynamic boundaries, and second, organic matter is in general not

distributed homogeneously in sediments, which can give rise, for example, to reducing

microenvironments in the aerobic oxidation zone.

Aerobic oxidation involves a single enzyme-mediated transfer of four electrons as O, is

reduced directþ to water; thus oxygen is considered to be a strong oxidant in aqueous

systems (Stumm and Morgan, 1981). It is typical of aerobes that they carry out, within

each cell, a complete oxidation to CO, of the organic compounds that they assimilate

(Jorgensen,l983). Due to their oxygen consumption, the oxic zone constitutes only a

rather thin layer in most cases. In terms of the amount of carbon oxidized, however,

oxygen is the principal oxidant in most lacustrine sediments (Fenchel and Blackburn,

1979). Oxygen is consumed not only by the mineralization of organic matter but also by

the oxidation of Fet*, Mn2* and other reduced substances, including sulphide species

and methane, which diffuse up into the oxic zone (Gobeil et al., 1987).

3-3

2-

2+Mn

O2

3o

f

H4

R ELATIVECONCENTRATION BIOGEOCHEMICAL ZONE

AerobicOxidolion

Denilrif icotion

Mongonese ond lron

Oxyhydroxide Reduclion

:Et-o-tl-Jo

Sulphote

Reduc f ion

Methonogenesis

Figure 3-1 Schematic dlstribution of biogeochemlcally ¡mportant specles in interstitlal waters ln ssdlrnents

showing the zonatíon tyficafff oUserved ln rnarine and lacustrine sedimcr¡ts. Thls dlsltibtflon wllsrretch;r compress iñ'respänse to changing physlcal, chcmical and blologlcal condhlons as

dlscussed in ths tsxt.

(@

SUBAOUEOUS DEPOSITION

Nitrate and the intermediate species nitrite, thermodynamically the first of the

secondary oxidants, are producecl during nitrification in the aerobic zone and consumed

by ctenitrification as o, becomes essentially depleted (smith et al., 1983). The nitrate

maximum sometimes observed at shallow depths in sediments reflects oxidation of

upward-diffusing NHr*; first to NOi by Nitrosomonas spp. and then to NO; by

Nitrobacter spp., in addition to the release to solution by aerobes of oxidized organic

N. Nitrate (and, for that matter, nitrite) will diffuse downward from the maximum to be

reduced by denitrifying bacteria below the oxic zoîe. Despite the relatively large free

energy yield associated with reduction.of NOj (Table 3-1), nitrate respiration as a

fraction of the total respiration in sediments is quantitatively minor which reflects the

low NO; concentration characteristic of bottom and pore waters (Sorensen and

Jorgensen, 1987).

The utilization by bacteria of the next-favoured electron acceptors, Mn and Fe oxides,

results in the dissolution of the solid phases and the release to pore water of dissolved

Mn2* and Fe2*. In practice, manganese oxide reduction commences slightly before

complete depletion of NO;, as shown in Figure 3-1. Fe2* typically appears in pore

water at a slightly greater depth than manganese, which reflects the different free

energy yields (Table 3-1) upon reduction of the oxides and the slower kinetics of

oxidation of Mn2* (Jacobs et al., 1985; Pedersen et al., 1936). Because the reduction of

MnO, commences in practice almost as soon as O, has been depleted, the depth at

which Mn2* first appears in pore water is a reasonably precise indicator of the sub-

bottom oxic-anoxic boundary.

Sulphate-reducing bacteria will consume SOo2- below the zone of iron oxyhydroxide

reduction, yielding HrS (which occurs mostly as dissolved HS- at the neutral or slightly

acid pH of most lacustrine pore waters). In sulphate-bearing lake waters, such as the

saline lakes of the B.C. interior (e.g. Hall and Northcote, 1986), sulphate reduction

probably accounts for a considerable portion of the total amount of organic material

oxidized in the sediments. In other meromictic lakes, such as those where post-glacial

isostatic rebound trapped seawater behind newly-emerged sills (e.g. Powell and

Sakinaw lakes in B.C. (see Sanderson et al., 1986), Garrow l-ake on Little Cornwallis

Island (Ouellet and Page, 1988; Bohn et al., 1981; Kuit and Gowans, 1981; Kuit, 1982)

and a number of lakes along the Nonregian coast (Strom, 1957)), the reduction of

sulphate at depth in the stratified water columns generates large concentrations of HrS.

Ferric iron (as FeOOH, for example) can be reduced by upward-diffusing sulphide

3-5


species either chemically or via bacterial catalysis. Evolved Fe2* will be rapidly

precipitated in the presence of dissolved sulphide as FeS of FeS' as discussed below.

Such authigenesis is responsible for the decrease of both sulphide and dissolved iron in

the sulphate-reduction zone in the Figure 3-1 schematic.

Reduction of carbon dioxide with concomitant production of methane proceeds once

sulphate has been depleted. In marine or sulphate-rich lacustrine sediments, the

inventory of SQ2- relative to the pool of degradable organic matter is large and the

sulphate reduction zone often thick. These phenomena, coupled with the fact that

sulphate-reducing bacteria are capable of oxidizing most types of organic matter

(except the most refractory compounds), usually preclude the onset of methanogenesis

in sulphate-replete deposits (Jorgensen, 7982). However, most fresh waters are

relatively depleted in SOo2-; a condition which commonly fosters methane production at

depth in lacustrine sediments rich in labile organic matter.

Studies of the hyclr<lgen ion activity in anoxic lacustrine pore waters in lakes of normal

(i.e. non-alkaline) hardness indicate that the pH falls typically in the relatively neutral

range 6.0 to 7.5 (Emerson, 1976; Carignan and Nriagu, 1985), even in cases where the

nverlvins lake water mav be aeid IBH's as low as 4.5, Carignan and Nriagu, 1985). The_ _ _--_* \r _

near neutrality of the interstitial waters reflects buffering in the anaerobic zone by

production of HCO; (alkalinity, via such reactions as sulphate reduction, iron

oxyhydroxide reduction, and FeS precipitation, which can be expressed collectively

(Carignan and Nriagu, 1985) as 4Fe(OH)¡(s) + +SOo'(aÐ + 9CH2O = 4FeS(s) + CO,

+ SHCO,- + 11 H2O), and NHo+ (via hydrolysis of ammonia, NHJ. Stumm and

Morgan (1981) ancl Morel (19S3) note that other reactions, such as the weathering of

silicates, as well as the ion exchange of H+ for other adsorbed cations, may contribute

to the buffering capacity of pore waters. Aerobic diagenesis in the oxíc zone can sustain

lower pH levels as a consequence of CO, production, but because this stratum is usually

thin in most lake sediments, the mean pH in lacustrine pore waters can be considered

to be more neutral.

3.2 Factors Affecting the Biogeochem¡cal Rubber Band

The intensity of sedimentary diagenesis is a function of two key variables, the

availability of oxidants, which is controlled both by physical and chemical factors, and

the demand for oxidants that accrues from the presence of labile organic substrate,

3-6


which in turn is controlled by the relative input of inorganic and organic phases and the

composition or "quality" of the organic compounds.

The depth to the oxic-anoxic interface in lacustrine deposits is controlled in part by the

rate atwhich oxygen can diffuse into the sediments (see Section 3.4.3) to replace that

consumed by bacteria and by oxidation of upward-diffusing reduced species. An

increase in the O, concentration in bottom water which will enhance the downward

diffusive flux of oxygen, will promulgate a deepening of the sedimentary redoxcline, all

other factors being equal.

The depth to the redoxcline also reflects the content of labile organic matter, which in

turn is controlled by the relative input fluxes of particulate organic and lithogenic

materials. The flux of inorganic detritus plays two opposing roles in this regard. First,

such material acts as a diluent, reducing the percentage of organic carbon in the

sediments, thereby reducing the bacterial oxygen demand per unit volume. Second, a

large input flux increases the linear sedimentation or burial rate of the organic fraction,

which more quickly removes a specific horizon from diffusive communication with the

overlying water, the source of most of the oxidant pool. It should be noted that the

concentration of dissolved oxygen in natural waters is low, rarely exceeding

400 ¡.tmol Ll. Therefore, in rapidly-accumulating deposits, the small quantity of orygen

in the interstitial water (which is also being quickly buried) is rapidly exhausted.

Beca¡se the consumect oxygen cannot be replaccd by cliffusion acting over a steadily

increasing distance, a high sedimentation rate promotes the establishment of anoxic

conditions at a relatively shallow depth.

The composition of sedimentary organic matter provides another major influence on

the rate of oxidant consumption during diagenesis. Detritus from vascular plants,

particularly cellulose (polysaccharides) and lignins (phenolic polymers), is particularly

resistant to bacterial degradation because of the high degree of crosslinking between

structural units (Emerson and Hedges, 1988). A low oxidant demand is associated with

the deposition of such materials. In contrast, carbohydrates and nitrogenous materials

derived from relatively protein-rich detritus such as planktonic matter are much more

readily degraded by microbes. Thus, input of algal remains to sediments will foster a

higher oxidant demand and shallower oxic-anoxic boundary than settled or buried

woocly clebris. It follows that highly productive rather than oligotrophic lakes will tend

3-7


to be floored by sediments which are anoxic at shallow depths, even given a higher

input of relatively refractory vascular plant debris to the latter.

Bioturbation, the mixing of sediments by burrowing animals, can increase signifîcantly

the depth of penetration of oxygen, both by advection as animals circulate bottom water

though their burrows, and by facilitating diffusion throughout the mixed zone and

particularly at its base. Three principal factors determine the extent of burrowing

activity: the organic matter content (i.e. the amount of food for the burrowers), the

oxygen content in bottom water, and the depth at which HrS accumulates in pore water.

As discussed earlier, these three variables are closely related, and must be considered in

concert rather than in isolation. HrS is toxic to animals and its presence limits their

activity. Thus, an ecological conflict exists: HrS, which is deleterious to bioturbators, is

generated at relatively shallow depths in organic-rich sediments, but the high organic

content of the same deposits tends to support an active infaunal community.

Bioturbation is bein g recognized increasingly as a very important influence on fluxes

across the sediment-water interface in both marine and lacustrine environments

(e.g. Kadko and Heath, 1934). In the absence of microstructure, and under steady-state

conditions, the cycling of redox-sensitive elements can be greatly enhanced by

bioturbation (Westerlund et al., 1986), although fluxes due to molecular diffusion are

the most important transport mechanism in the upper few centimetres in cases where

concentration gradients are steep. Advection associated with bioturbation is more

important as a transport mechanism for sediments below this zone (Emerson et al.,

1e84).

Bioturbation can alter concentration gradients in pore waters by increasing the effective

surface area of sediment particles exposed to adsorption-desorption and precipitation-

dissolution reactions (Kadko et al., 1987). Bioturbators which probe the upper reaches

of the suboxic or anoxic zones in sediments can promote the oxidation of authigenic

sulphides, which fosters release of metals to interstitial solution, in effect creating a new

dissolved metal source at depth (Emerson et al., 1984).

Assessment of diagenesis in bioturbated sediments is further complicated by:

. differing reaction rates around burrows resulting from concentration gradients

induced by adding a third dimension to diffusion;

. temporal variability in burrow location;

3-8


burrow size variations; and

physical discontinuities associated with burrow structures (Aller, 1983)'

In summary, the presence of an active infauna in sediments tends to thicken the aerobic

zone and promote enhanced fluxes of some dissolved constituents in both directions

across the sediment-water interface.

It has been demonstrated in a number of studies ttrat rates of diagenesis and

consequent stretching or compression of the biogeochemical rubber band vary

seasonally in lacustrine and marine sediments in temperate and northern latitudes (see,

for example, Jorgens en, 1977; Holdren and Armstrong, 1980; Sholkovitz and Copland,

l9g2; and Klump and Martens, 1981). A number of factors control such variations.

First, dissolved oxygen contents in bottom water reflect the influence of stratification,

which has a strong seasonal character in most lakes. Second, primary productivity

varies seasonally, typically being highest during the well-illuminated spring and summer

months; the settling flux of organic matter directly and almost immediately reflects such

variability. Third, seasonal warming of sediments and overlying water promotes

increased rates of bacterial metabolism, which amplify the benthic oxidant demand.

The net effect of these influences is to compress the biogeochemical zonation

(i.e inrensify diagenesis) during the late spring/summer period. Associated steepening

of concentration gradients promotes larger benthic fluxes of most constituents involved

in near-surface diagenetic reactions.

3.3 Diagenetic Consequences

3.3.1 Role of Oxyhydroxide Phases

Poorly ordered, often amorphous, iron and manganese oxides and hydroxides are

ubiquitous constituents in lacustrine sediments, frequently occurring in concentrations

rangingfrom about 1to >10wt. Vo (e.g.Hamilton-Taylor, 7979; Farmer et al', 1980;

Carignan and Nriagu, L985; Cornwell, 19S6). These phases are important diagenetically

because of their participation in redox reactions as discussed above, and because their

surfaces provide sites for the removal of trace metals from natural waters. Kadko et al.

(19S7) note that sedimentary iron and manganese oxides are quantitatively very

important scavengers of trace metals because of their high specific surface areas, high

negative surface charge, and high cation adsorption capacity over the pH range of most

a

3-9

SUBAQUEOUS DEPOS¡TION

natural waters. A number of studies have addressed the specific adsorption of trace

metals on Mn and Fe oryhydroxides (e.g. Balistrieri and Murray,1982,1984, 1986), and

it is clear from this work that the distributions in sediments of a range of trace metals

including Cu, Ni, Zn, Cd, Pb and Mo are strongly influenced by the behaviour of the

oxide phases. In many cases, sedimentary concentration profiles of trace metals which

decrease with depth parallel those of the major oxides (e.g. Cornwell, 1986). Such

distributions do not necessarily reflect recent anthropogenic influences but may instead

be a reflection of a natural diagenetically-produced oxide-trace metal association.

Iron and manganese oxides are characteristically recycled during early diagenesis in

sediments (see for example Pedersen et al., 1986, and Carignan and Nriagu, 1985).

Reductive dissolution below the aerobic zone releases Mn2* and Fe2* to pore water'

which supports an upward diffusive flux of both ions. Oxides reprecipitate when the

dissolved species encounter O, in the aerobic zone, producing a stratum relatively

enriched in manganese and iron. Under steacly-state conditions, this layer is

subsequently buried and the cycle is repeated. This recycling procedure usually

maintains solid-phase Mn and Fe enrichments near the sediment-water interface

(Figure 3-2). The thickness of this oxide-rich zone is, to a first approximation, inversely

proportional to the intensity of diagenesis in a given sediment columr¡ and can

therefore be used as a rough indicator of the comparable diagenetic status of sediments

from different locations.

3.3.2 Authigenic Sulphides

Under anoxic conditions, hydrous oxides and oryhydroxides are replaced by sulphides

as the dominanr solid authigenic phases (see Section 3.4.1)" The log solubility product

(i.e. p\os) ranges from about 25 (ZnS) to about 53 (HgS), which implies that a large

proportion of trace metals dissolvecl in anoxic, sulphicle-bearing pore waters should

precipitate and be fixed in the sediments as solid sulphide phases (Framson and Iæckie,

1978).

Dissolved iron precipitates readily as metastable FeS in the presence of HrS. This

monosulphide phase is common in sediments, and because it is known to form solid

solutions (Framson and læckie, 1978), it is suspected that other metals aÍe

coprecipitated by FeS (Gobeil et al., 19S7). The solid phase activity of a trace metal in

3-10

l+ Mn (s)

FeoOH(S)

FeS (S)

_2+FE

RELATIVE CONCENTRATION

Tt--o-t¡Jo

Figure 3-2 Schernatlc dlstrlbutlon wlth dåpth <ú rnanganese and lron specles ln lnterst¡tlal water and solld

sediments Steady state b assi¡rned- The near-surf,ace oxyhydroxHe enrlchrn€nts reflect lnput of

deuital oddes'to-ûþ .sedlmer¡t surfacs, contln¡rcus ¿SsAutnn at dsptlt up¡rard dltruslon of

dlssofued specles, ard'reorecloÈatbn lrthe aeroblc ztxì€-

(@

SUBAQUEOUS DEPOSIT¡ON

a solid solution is a function of both its solid phase activity coefficient and its mole

fraction (Jacobs and Emerson, 1982), and is usually < L. This decreased solid-phase

activity is especially important in governing the solubility of the minor component of the

solid solution. Stumm and Morgan (1981) note that the observed occurrence of metals

in sediments formed from solutions that appear to be formally unsaturated (ignoring

solid solution formation) with respect to the impurity can often be explained by solid

solution formation. However, Jacobs et al. (1985) used carefully measured field data as

a basis for suggesting that chemical equilibrium with an impure FeS phase is not the

major process controlling the dissolved trace metal concentration in sulphidic waters;

instead they invoked the formation of pure, specific metal sulphide phases as the

probable leading control.

Srite forms in anoxic sediments following the reaction of FeS with elemental sulphur

(Berner, 1964, and Rickard, 1969; equations (1) and (2)), or of ferrous iron with

elemental sulphur or polysulphide ions in the presence of dissolved sulphide species

(Rickard, L975, and Howarth, 1979; equations (3) and (4))' viz:

Fe2*+HS-=FeS+H*FeS+So=FeSzFe2* + S*" * HS- = FeS, + S*-, + H+

Fe2*+So+HrS-FeSr+ZfI+

Davies-Colley et al. (19S5) note that free sulphides and polysulphides can react more

directly through bacterially catalyzed reactions with Fe and Mn oxides to produce

poorly ordered, black, fine-grained monosulphides via

2FeOOH + 3HrS = 2FeS + So + 4}J.2O (5)

The monosulphide phase may then react with elemental sulphur to form FeS, as in (2)

above. Hence, sulphides that diffuse into the oxide reduction zone may be oxidized by

FeOOH (or MnOr) to form S0 and pyrite. Because of its ability to coprecipitate trace

metals, pyrite is thought to be a significant sink for metals in anoxic sediments (Dyrssen,

1e8s).

(1)()\(3)

(4)

3-12


3.3.3 Benthic Fluxes

As noted earlier, oxygen is the primary oxidant in lacustrine systems, in part because

sulphate occurs only in small concentrations in fresh waters, unlike the case in seawater'

Dissolved O, in bottom water rvill diffuse into sediments along the gradient of

decreasing concentration established by consumption at or below the sediment-water

interface. The quantity of oxygen (or other oxidants) which can diffuse per unit time to

a given depth below the interface (Jr, in ¡rmol cm-2 secl) is governed by the steepness of

the concentration gradient, dcldz (in pmol .-t), sediment porosity (4, dimension-

less), tortuosity (a measure of the tortuous path length followed by a diffusing molecule'

determined by measuring the electrical resistivity of wet sediments and described by a

dimensionless variable, F, the formation factor (Manheim, 1970)), and the molecular

diffusion coefficient D,,o,rro which varies as a function of temperature and pressure and

is given in units of .-ã'-íe.-t. These variables collectively form Fick's First l-aw of

diffusion, viz:

J" = -(D¡cr, rro /r)o{ac/ az¡

In practice, estimation of the concentration gradient is complicated by the presence of a

thin, essentially stagnant layer immediately above the interface (the diffusive boundary

layer) in which eddy diffusivity is zero and solutes move only via molecular diffusion'

This zone is typically a fraction of a millimetre to several millimetres thick (Boudreau

and Guinasso, 1982). The difficulty inherent in estimating this thickness accurately

limits the precision of all flux calculations where concentration gradients are very steep'

such as in sediments where anoxic conditions prevail within several millimetres of the

interface.

Although oxidants will, in general, diffuse into sediments, other constituents, including

aqueous species regenerated from organic nitrogen and phosphorus and dissolved

metals, wilt diffuse upward if concentrations in shallow pore water are higher than in

overþing bottom water. Such benthic effluxes are quantitatively important in the

cycling of nutrients in many lakes (Holdren and Armstrong, 1930)' Dissolved metals

appear to diffuse readily into lacustrine sediments where anoxic conditions occur at

shallow depths (e.g. Carignan and Nriagu, 1935). This phenomenon can be explained

by the precipitation of mineral phases such as ZnS, NiS and CuS within the upper 1-

2 cm of the sediments. Such downward fluxes and precipitation reactions are believed

3-13


to account for a substantial portion (on the order of.50Vo) of the accumulation of Ni

and Cu in the sediments of polluted shield lakes near Sudbury, Ontario (Carignan and

Nriagu, 1985) and of Zn in an acid lake in southeastern Quebec (Carignan, 1985). In

contrast, dissolved metals concentrations in pore waters immediately below the

sediment-water interface of unpolluted coastal marine sediments are t¡pically higher

than in the overþing bottom waters. In these cases, the sediments act as a source rather

than a sink for metals (e.g. Pedersen, 1985; Westerlund et al., 1986), and reflect the

release to solution of organically-bound metals as organíc matter near or at the

interface is aerobically degraded. Diffusion from a near-surface mar<imum

characteristically occurs in both upward and downward directions; the latter reflecting

low dissolved metals concentrations in the sulphidiczone at depth.

Sundby et al. (1986) and Westerlund et al. (1986) carried out a series of experiments on

benthic fluxes of metals which illuminate particularly well the role of orygen in

governing diagenetic release near the sediment-water interface. Although the

experiments were performed on shallow (ó m), organic-rich, marine sediments, the

results are fully applicable to lakes. These authors used stirred benthic chambers in

which the pH was kept constant by the addition of NaOH and the oxygen content was

either kept constant by the addition of O, via a capillary or was allowed to be depleted

by benthic respiration. Under orygenated conditions, which probably extended to only

a few millimetres depth, Cd, Cu, Zn and Ni were released to the overþing water while

Co, Mn and Fe were taken up. When the O, concentration in the water was allowed to

fall to zero, dissolved Cd, Cu, Zn and Ni concentrations in the chamber fell

significantly, indicating uptake by the sediment, and Co, Mn and Fe were released from

the sediment surface. Three main reactions appeared to be responsible for the

observed behaviour. First, oxidative degradation of organic matter at the interface

releases associated trace metals to solution (probably much of the Cd, Co, Cu, Zn and

Ni). Second, in the orygenated chamber, the high concentrations of Mn2* and Fe2*

initially present were oxidized to form particulate oxyhydroxides whose occurrence was

apparently restricted to the top millimetre of the sediments. The precipitation

behaviour in this experiment of both of these elements but particularly iron is consistent

with the rapid kinetics of oxidation at neutral pH (Stumm and Morgan, 1981). Cobalt is

readily scavenged from solution by FeOOH, which e4plains its observed covariance

with Fe2*. Although the other metals would also be scavenged by the oryhydroxides,

their lack of depletion in the chamber water while oxygen was present probably

reflected the dorninarìce of addition from the presumed degrading organic source.

3-14


Third, the decrease of the trace metals concentration when oxygen became deplete was

ascribed to their precipitation at very shallow depths as discrete sulphide phases. The

concurrent release of Mn, Fe and Co was attributed to reductive dissolution of

oxyhydroxides as the redox potential in the upper millimetre or so of the sediments fell.

These observations conform with the behaviour predicted by the biogeochemical

zonation theory discussed earlier, and demonstrate quite clearly that the presence or

absence of oxygen is a fundamental determinant of the behaviour and distribution of

dissolved metal species in aqueous systems.

3.4 Application to submerged sulphlde-Bearing Mine waste Deposits

The theoretical framework of early diagenesis as it pertains to the behaviour of metals

in natural sediments, described on the preceding pages, is equally applicable to

submerged mine tailings deposits. In this section we consider the role that diagenesis

should ptay in governing the mobility of metals in sulphide-bearing lacustrine tailings

and mine wastes.

3.4.L The Oxidation Problem: TheotX

As noted earlier, sulphide minerals are invariably only very sparsely soluble, and under

anoxic conditions, they can be considered to be stable phases. In the presence of

molecular oxygen, however, this stability is greatly reduced, to the extent that some

sulphide minerals (e.g. FeS) are readily chemically (i.e. abiotically) oxidized. Such

reactions yield a range of alteration products which are typically amorphous

oxyhydroxide phases, carbonates and sulphates with solubilities considerably higher

than those of their precursor sulphides. Solubility products for selected monosulphide

minerals in freshwater (ioníc strength, I, of zero) are listed for reference in

Table 3-2 atong with equilibrium solubility constants for oxides and hydroxides,

carbonates and hydroxide carbonates. Note that it is not possible to compare the

solubility constants of sulphides with the other phases directly, given the differing

reaction stoichiometries. It is more instructive to compare the concentration of the free

ions of the dissolved metals in solutions which are in equilibrium with the solid phases

of interest, as in Figure 3-3. Note that at a pH of 6, and ignoring the formation of

hydroxo metal complexes, the concentrations of Cdz*, Ztf* and Cu2* are quite high,

while that of dissolved Fe3* is extremely low; this observation reinforces the fact that

iron oxide phases are extremely insoluble in natural orygenated waters, in contrast to

3-15


fable 3-2

Gonstants for Selected Solublllty Equlllbrla

Sulphides log K,25"G,I = Q

MnS(s) = Mn2++52-

FeS(s) = Fe2*+S2'

ZnS(s) = Zn2*+S2-

CdS(s) = Cd2* +S2'

CuS(s) = CuZ*+S2-

PbS(s) = Pb2++S2'

HgS(s) = Hg2*+S2'

-13.5

-18.1

-24.7

-27.O

-36.1

-27.5

-52.7

Oxides and Hydroxides

c-FeOOH(s) + 3H+ = Fe3* + 2H"O

(am) FeOOH(s) +3H+=Fe3* +ZH"O

ZnO+ 2H+ =Zn2+ +zHzO(am) Zn(OH)r+ zH+ = Zn2+ + 2HrO

CuO(s) + 2H+ = Cu2* + HrO

*K**K.o

*K**K.o

*K*

= 0.5

= 2.5

= 11.14

= 12.45

Carbonates and Hydroxide Carbonates

Zn(OH).,.r(CO.)o.o(s) + 2H+ =Zn2+ + 1.6H2o + o.ccOr(o)

ZnCO.(s) +2H+ =zn2* +Hro+Co2(g)

Cu(OH)(CO3)o.r(s) + 2 H+ = Cu2* + 3/2HrO + t /zCQr(9|PbCO3(s)=Pb2+ rCO.z'

COCO.(s) + 2 H+ = Cd2+ + HrO + CO,

trinOO.(s) - Mn2+ + CO.2'

*Koso = 9'8*Kpro = 7'95*Koso = 6'49

K¡o = -13'1

*Kpro = 6'44

K.o = -10'4

SourcE: Stumm and Morgan (19s1). Asterlsked constant3 reproscnt_tho orlglnal lgrmlno¡ogJ ln tlr. ¡blct ol- õon¡un6 publlsñed ¡li ué. S¡tàn and ÀC. Uirt"li, Su¡¡nty Con¡tant-¡ of Metat'lon êomplorcr, Spcclal

Publlcallons, Nos. 17 and 25, Chemlcal Soclety, London, 1964 and 1971.

3-f6

2r

2+ Ag"zn2'2+

Co

+Fe M9Cu 2*

Fe 3* Al3*

2

¿-aNoUgìoI

3

4

5

66 I ro t2? 4

pH

Figure 33 Solubiltty of selected oxides and hydroxHes plotted as free metal lon concsntradon ln equillbrlum

with solH o<iJes or hydroxkles vs. pH. Noto that the plot is based on I - 0 and that the

occurencE of hydroxo meal complo<es has not been taken ¡nto accounL After Stumm and

Morgan (1981).

@


most other metal oxides. As pH decreases, solubility of all the solid oryhydroxide

phases increases; a drop of one pH unit promotes a rise of roughly two orders of

magnitude in the concentration of dissolved metals. Similar considerations apply to

metal carbonate equilibria in non-alkaline solution, and the solubility of hydroxide-

carbonates, such as hydrozincite (Figure 3-4).

A simple comparison illustrates the solubility contrast between sulphides and their

oxidation products. Consider, for example, ZnS and hydrozincite (Zn (OH).(CO3)2).

Ignoring complex formation, and using solubility constants from Stumm and Morgan

(1981) and the specific conditions pH = 6o I = l0-2'a, T = 25'C, total dissolved

CO, = 10-35 M and P = 1 atm., then lZrf+l in equilibrium with hydrozincite will be

appro*imately 106 M. In a similar but anoxic solution, with [total HrS] = 106 M (a

lower concentration than is typically encountered in anoxic pore waters), a pKrn for

sphalerite of 1L.4 (based on an equation of the form (lZtf+][HS])/[H*], ut formulated

by Jacobs et al., 1985), the concentration of. Zrf* in equilibrium with ZnS will be about

10-12 M, some six orders of magnitude less than that for hydrozincite. Clearly, the

presence of sulphide in solution will establish extremely low dissolved metal

concentrations in most cases; unfortunately, much higher metal levels will characterize

oxygenated waters in contact with sulphide oxidation products.

3.4.2 The Oxidation Problem: Practice

In general, oxidation in natural sediments is inhibited where rapid accumulation is

accompanied by a high organic matter content. As noted above, because the

concentration of dissolved O, in water is rather low, the available oxygen in pore waters

will be rapidly depleted given a respiring bacterial communiry and a linear

sedimentation rate high enough to remove newly deposited sediments reasonably

quickly from diffusive communication wíth overlying bottom water" Thus, the

combination of a high organic load and rapid sedimentation will ensure the

establishment of anoxic conditions at shallow depths and prevent the oxidation of

deposited or authigenic sulphide minerals.

One other major factor which bears on the potential for oxidation of tailings is grain

size. Because it is desirable to limit the surface area of sulphide particles exposed to

oxygen during and after deposition in lakes, a mean particle size as large as possible is

preferred; however, this must be balanced by metallurgical feasibility"

3-18

o

-4

o 2M

ECO,=¡9 - 4

M

*ñN cNIt-,

Ql)o

-12

-t6

46 to t4

PH

Figure 34 The concer¡t¡atlon oî Tstz+ in equillbrium with hydrozinc¡te vs. pH (25XC, I - 4 x 10€, pKs -64. Note that hydroz¡ncite is not stable 6/er tho entlre pH range shown, and that hydroxo Zncomplexes have been ignored ln the calculatlon of the curves (after Stumm and Morgan, 1981).

After Stumm and Morgan (1981).

(@


Oxidation of submerged tailings and mine wastes and release of contained metals would

be entirely prevented if such materials could be deposited in lakes or fjords which

contained anoxic bottom water; such environments are relatively rare, however. In

orygenated basins, it is clear that rapid vertical accumulation of tailings or waste rock is

preferable to lateral dispersion given the need to limit the flux of orygen into the

deposits. Similarly, the admixture of organic matter, derived either from codeposition

of natural organic detritus, or by purposeful addition, will promote desired rapid oxygen

consumption.

Most lakes are floored by sediments which are naturally anoxic at shallow depths.

Thus, following cessation of tailings or waste rock discharge, it is a reasonable

expectation that the deposits will be covered by an accumulating veneer of natural

sediments which will act as a permanent protective barrier to oxidation. The time that

elapses between the cessation of discharge and the development of an oxic-anoxic

boundary in the accumulating cover layer will depend on the local character of

sediment deposition in the recipient basin.

3.4.3 Speciation of Metals in Natural Waters

Speciation refers to the distribution of a trace element over the suite of complexing

inorganic and organic ligands present in natural waters. Understanding impacts on

biota requires consideration of this phenomenon given the relatively recent recognition

that it is not the total amount of a metal pollutant in aquatic systems that is most

important but rather the concentration of biologically available species of the metal

(e.g" Sunda and Guillañ,1976; I-ewis and Cave, 1982).

Significant ligands for trace metals in aerobic waters include OH-, Cf, CO32-, HCOj'

SO42-, organic molecules and macromolecules, surface sites, and to a lesser extent'

phosphorus and silicate species and NO; (Bourg, 19SS). Reduced or intermediate

sulphur species are additionally important in anaerobic environments. In freshwaters,

in which the concentration. of inorganic ligands is quite low, the complexing capacity

varies quite widely, largely as a function of the concentration of dissolved organic

matter and the presence of particulate phases. Complexed metals can exist in true

solution, in association with colloids, or adsorbed onto particle surfaces (Figure 3-5).

The complexing capacity of sediments is relatively high because of the enriched content

of dissolved organic carbon compouncls (largely humic and fulvic acids) in pore waters

3 -20

IRONMANGANESE

IGH MOLECULAR WEIGHTORGANICS

INORGANIC LIGANDSLOW MOLECULAR WEIG

ORGANIC LIGANDS

Cootings Sorption Sorplion Sorption Sorplion

SedimentotÍonResuspension

DEPOSITED SEDIMENTS

Ageing effecls, redistribution of lroce melols over sedimentory phoses

Releose of components lo the interslitiol wolers

Figure 3-5 Sr.rmrnary of rnalor processes and mechanlsms ln the lnteracllons betwsen dlssolved and sol'vjmetals specles ln surface waterE

SUSPENDED MATTER( porlÍculole metol speciotion )

METALCOLLOID

COMPLEXES

METAL( IN )ORGANICCOMPLE XES

METALS¡+

Source: Salon¡onc and Forgner (19S4).

@


and the presence of abundant particle surfaces. Stumm and Morgan (1981), Solomons

and Forstner 1ílA+¡, and Bourg (1988) describe speciation models for metals in natural

waters which take into account the relative affinities of ligands in solution and on

surfaces. One example is shown in Table 3-3. Note in this model that for metals such

as Cu, there is considerable contrast between the complexing ability of citrate (high)

and glutamate (low), and a widely different degree of complexation of citrate of Cu

compared to Ag. Note also that complexes of metals with inorganic ligands are in

general quantitatively unimportant in the fresh water model because the major cations

occur in low concentrations; only COr" plays a significant role (Table 3-3). The

addition of a relatively strongly-complexing organíc ligand such as citrate reduces the

proportion of Cu complexed by COra by a factor of 500 (10-s'z M versus 10-7 M).

Clearly, speciation in interstitial and lacustrine waters will be a complex function of the

concentration and type of inorganic and organic ligands and particulate matter present,

pH, total metal concentration, the redox state, and the flow rate of solution through the

aquatic system (Bourg, 1988). This complexity has been addressed in previous and

extant computer models (e.g. Mantoura et al., L978; Turner et al., 1981), and in

laboratory studies of metal partitioning in freshwater sediments (Tessier et al., 1979,

1982; Rapin et al., 19S6). The empirical studies have highlighted the inherent

variability that is characteristic of metal speciation and partitioning in natural

sediments. It is not yet clear, however, to what extent the results of these predictive and

empirical studies can be applied in assessments of impacts of pollutant metals on biota;

much work remains to be done in this area.

3 -22

Table 3-3

Equilibrium Model: Etfect of Complex Formation on Distribution of Metals in Aerobic Waters

" All concenlraüons arc alven 13 - log (mol lttre'l). Charges of sp€cles are omltted.b

Organlc matter ol epproxlmately composldon C--H--O--N.

" Total conccntr¡üon of mctal ¡pscle¡. n. *nJiitLtbIr or rre.ry m€tat! glyon arc hlghcr tùan fose typlcatly found ln unpolluted ¡olwater or ficrh wltqr' Thc ¡ehthr oflectl

of complcr form!üon tr trre,. olomcnt¡ rro tndgpondont ol th. total ".ncentraüon

ol thelc clementl; water¡ are ln equlllbrlum wttñ fe(OH).(r)'d Tho conccnlraüonr tclo? to tñc ¡um of ¡ll complexes, for example, CuCtt, CuHClt CuCtt .

t A desh motns tñat no dlblltty conttants are avallablc fo? ruch complexes.r P.tc"ntage of cach llgand bound to metal lon¡.

2.72

3.72

17.73

7.u9.93

6.72

7.76

8.04

9.19

CaHCO3,4.6

M9SO.,5.1

Fe(OH).,8.7

MnSO.,8.5

CuCO.,9.7

ZnSO1,8.2

cdso4,9.2

Pbco3,7.1

Agq,9.5

7.O

8.0

19.0

11.3

13.1

10.3

1 1.5

11.0

13.8

f.5

5.2

7.O

7.2

9.7

7.O

10.5

9.2

8.9

17.5

98.2

9.1

8.0

15.1

11.5

9.4

9.6

11.5

10.1

13,3

0.3

5.7

29.035.2

1

%l

11.1

9.4

8.6

':'

0.4

1.3

8.9

9.6

8.8

11.4

9.1

9.7

9.2

8.6

9.1

5.6

7.1

lnorgan¡c fresh water plus 7 x 106 mol litre-r of each of the indicated organic ligands

corre.ponãing to 2.g mg titrà'1 of soluble organic carbonb. lnorganic ligands remain unchanged

MaiorlnorganicSpecies

Organic Complexesd'"

Free M

Phthalate5.30

Glutamate5.16

Gl¡æinate5.16

Tartrate5.34

Citrate6.91

Acetate5.16

FreeLigand

C,a

Mg

Fe(llD

Mn(ll)

Cu(ll)

z'(lD

cd(lD

Pb(r)

As(l)

2.7

3.7

Satd.

7.O

7.O

6.7

7.7

7.O

9.0

2.72

3.72

17.70

7.U7.46

6.72

7.73

8.O2

9.19

CaHCO3,4.6

MgSO., s.1

Fe(OH).,8.7

MnSOo,8.5

CuCO.,7.2

ZnSO4,8.2

cdso4,9.2

tuco3,7.1AgCl,9.s

9'r

35

95

93

9s

95

v.low

9.5

65

F¡esh Waterlnorganic lresh weter, PH = 7.O,É" C, lree ligands:pSO¿ 3.4; pHGO.3.1; PCO.6.1; PCI 3.3

Free M lonFree

M"T

M

MajorSpecies

After Stumm and Morgan (198f )

4.0 LIMNOLOGICALAND BIOLOGICAL CONSIDERATIONS

The preceding sections described contemporary understanding of the acid mine

drainage process and how subaqueous disposal might be used as a means of mitigating

the problems associated with it. This section discusses the potential biological impacts

of subaqueous disposal of reactive mine wastes in a freshwater environment. Impact on

macrobiota is addressed with respect to species specific effects of turbidity and

seclimentation on various biological parameters, toxicity, and the potential for

bioaccumulation and biomagnification. In addition, the implications of limnolory to

site selection are considered.

It is important to note that the biological impact of disposing mine wastes into

freshwater varies with the study organism, the physical and chemical nature of the

waste, and the limnological characteristics of the receiving waters. In addition, the

biota of these waters are not always the passive victims of deposited materials. Some

organisms re-work the sediments thereby altering the physical and chemical nature of

the microhabitat and its suitability for other life forms. These considerations are also

incorporated in the following discussion.

4.1 MicrobiologY

A search of the literature indicates that there is very little published material on the

microbiology of mining wastes following subaqueous deposition. In addition there is

little ecological information on the two groups of micro-organisms of primary interest in

such situations: the sulphide oxidizers and the sulphate reducers.

Karavaiko (1978) in his review of the microflora of land micro-environments, reports

that various authors have found that many and cliverse micro-organisms are present in

the products of rock degradation. Many require a source of carbon for growth (the

heterotrophs) while the autotrophic organisms obtain their energy for growth by the

oxidation of reduced forms on nitrogen, sulphur and iron.

Ore deposits, being a source of both reduced sulphur and iron compounds, are an

excellent ecological niche for members of the genus Thiobacíllus as well as similar types

of micro-organisms such as those cited by Karavaiko which can oxidize Sb*3, Co*3, and

As*3.

4-1

LIMNOLOGICAL AND BIOLOGICAL CONSIDERATIONS

4.1.L Sulphide Oxidizers

The most important organism relative to microbiological alteration of sulphide bearing

materials isThiobacillus fenooxidaru. This organism, which was discovered in the acidic

drainage from coal mines in Pennsylvania in 1947 (Colmer and Hinkle, 1947; Colmer et

al., 1950; Temple and Colmer, 1951), is characterized as an authotrophic, motile,

aerobic, gram negative rod. It usually measures 0.5 by 1.5-2.0 microns and is capable of

oxidizing ferrous iron, thiosulphate, sulphur and metallic sulphides (McGoran et al.,

1969) to obtain energy for growth while using oxygen as the final electron acceptor. It

uses carbon dioxide as its sole source of carbon and requires an environment with an

acidic pH.

The literature refers to two other organisms, Fenobacillus fenooxídans (L-eathen et al',

1956) and Fenobacillus suþoxidans (Kinsel, 1960), which have the same characteristics,

but most researchers consider them all to be the same organism (Hutchinson et al.,

1e66).

In this report, reference to T. ferrooxidans includes all the acidophilic thiobacilli capable

of oxidizing metallic sulphides and ferrous iron.

Other thiobacilli are frequently cited as being associated with the microbiological

leaching environment but they cannot oxidize ferrous iron. Thiobacillus thiooxidans has

very similar characteristics to T. fenooxídaru but it cannot oxidize iron nor can it oxidize

metallic sulphides with the exception of sodium sulphide (unz and Lundgren, 1961).

T. fenooxídøru has minimal requirements for growth. It obtains energy from the

oxidation of reduced sulphur and iron compounds. It requires CO, as a carbon source'

NHo* for the formation of cellular proteins, phosphate for the formation of adenosine

triphosphate (ATP), other trace nutrients for various cellutar components, and oxygen

as the ultimate electron acceptor. Oxygen and the various inorganic nutrients are

normally present in ample supply in the environment associated with terrestrial ore

bodies. Ammonia nitrogen may be limiting, although there is a preliminary report in

the literature that suggests the organism may be able to fix atmospheric nitrogen

(Mackintosh, 1971).

4-2

LIMNOLOG ICAL AND BIOLOGICAL CONSIDERATIONS

T. thíooxídøru has very similar characteristics except that suggests it only utilizes sulphur

or thiosulphate as its energy source. It cannot oxidize the metallic sulphides or iron

found in ore bodies (U* and Lundgren, 1961).

Both X fenooxidans and T. thiooxidatu are extremely acid tolerant; pH values of less

than t have been generated frequentþ in culture laboratory studies at B.C. Research.

The upper limit for sulphide oxidation by T. fenooxídans is around pH 4, whereas

sulphur oxidation has been found to occur at pH 5 (McGoran et al., 1969). The upper

limit for sulphur oxidation by T. thíooxidans is arouncl pH 7. The upper limit for the

oxiclation <lf ferrous iron by the former organism is difficult to determine since the

auto-oxidation of this iron is extremely rapid above pH 3.5. In general, the most

favourable pH range for growth is 2-3.

In spite of its acidophilic nature and the low pH range for oxidation of its two main

energy sources, it is known that I fenooxídans will survive for extended periods of time

at a pH of 8. Studies showed that the organism survived for 24 hours at pH 9, but no

viable organisms were found after 48 hours at that pH.

Most strains of. T. fenooxidans have an optimum temperature in the range 28-37'C. In

B.C. Research studies it would not survive at 45'C. The rate of activity of the organism

drops off as the temperature decreases below 28" C. Oxidation rates are very slow

below 5 "C.

As mentioned above, oxygen is the ultimate electron acceptor and the reduced oxygen

combines with the oxidized sulphur to yield sulphate. Two moles of oxygen are

required for every mole of sulphide oxidized, thus the rapid oxidation of sulphides

requires significant oxygen.

The other gas required by the bacteria is carbon dioxide. For all intents and purposes

this is the organism's sole source of carbon. It can assimilate other sources of carbon

but it is unclear if it is truly capable of growth using carbon compounds as an energ]

source. Normally, the availability of carbon dioxide will not be a growth limiting factor.

T. fenooxidans is not inhibited by high concentrations of most heary metals. In studies

at B.C. Research, concentrations as high as 62 g/L copper, 120 g/L zinc,26 g/L nickel

and 20 g/L arsenic have been tolerated. Two metals that have been shown to be toxic

are molybdenum (100 mg/L) ancl uranium (1100 mglL). In both cases, the toxicity

4-3


appears to be due to the fact that these metals occur as cations at the acidic pH values

preferred by the organism. Other anions toxic to T. fenooxidans aÍe chloride and

nitrate. Anions, with the exception of sulphate, are generally found to be inhibitory.

As shown by the above brief sumûâry, T. fenooxidaru occupies a rather unique

ecological niche. This niche is relatively easy for the organism to develop and maintain

in the terrestrial environment. Even in alkaline host rock, the organism can attach itself

to an exposed sulphide surface (as long as it is moist), obtain the orygen it needs and

initiate oxidation in that micro environment. The acid generated maintains the desired

conditions and then gradually neutralizes the surrounding alkaline rock thereby

expanding the size of the desirable environment. As more sulphide is exposed to the

acid environment, the process accelerates.

In the subaqueous environment represented by a lake bottom or a sea bed, however, it

is much more difficult, if not impossible, for the bacteria to establish a

micro-environment. The amount of oxygen available is severely reduced" On land, air

containing ZTVo oxygen is normally in contact with the thin film of moisture in which the

bacteria are located. under water, oxygen availability is inhibited by the limited

solubility of O, the depth of the water column and by competing oxidative processes.

Oxygen transfer thus becomes a rate-limiting step.

In addition, in the subaqueous environment the micro environment is not isolated but is

constantly subject to dilution by the mass of the surrounding water. Even if there were

no active currents which would cause rapid dilution, the diffusion of acid out of the

micro-environment would be more rapid than the diffusion of orygen into the

environment given the very high diffusion coefficient for protons. As stated above, it

takes two moles of oxygen (ignoring any oxygen required for iron oxidation) to produce

one mole of sulphuric acid (which is equivalent to two moles of hydrogen ions). It is

unlikely, therefore, that the concentration of hyclrogen ions will ever accumulate to a

level that causes a drop in pH sufficient to accelerate microbiological activity.

Another inhibitory factor in seawater is chloride ion toxicity. I-aboratory studies at

B.C. Research have shown that normal bacterial leaching of metals occurred when the

medium contained l07o seawater, reduced leaching occurre d at 257o seawater, but no

leaching was evident at 50Vo and 1007o seawater. Based on these results, it was

concluded that the normal terrestrial strains of T. fenooxídans are unlikely to oxidize

sulphide minerals when the salinity exceeds 13 ppt.

4-4

LIMNOLOG ICAL AND BIOLOGTCAL CONSIDERATIONS

To the best of our knowledge, strains of T. fenooxidaru have not been isolated from

seawater. Marine thiobacilli have been isolated, however, (Tilton et al., L967) and they

are considered to be distinctþ different from non-terrestrial types. In addition, several

of the strains isolated were capable of oxidizing HrS (Adair and Gundersorl' 1969).

4.L,2 Sulphate Reducers

The other group of micro-organisms associated with subaqueous disposal of mining

wastes that have a potential influence on the environment are the sulphate reducing

bacteria. These organisms are responsible for dissimilatory microbial sulphate

reduction, a process in which sulphate is used as the electron acceptor in the oxidation

of organic matter. This reduces sulphur from the + 6 oxidation state to sulphide

(-2 oxidation state). Although the Desuþvibrio are the most commonly studied

organisms of this type, Desulþtomaculurø also have this capability (Miller and Hughes,

1968).

Desutþvíbrio are small, gram negative, curved rods which are strict anaerobes. They

are motile by means of a single polar flagellum and grow ín both fresh and seawater at

natural pH values.

The interaction of Desulþvibnb with mining wastes is based on its ability to generate

sulphide ions which interact with metals to form insoluble sulphides. It requires

significant quantities of organic matter to drive its metabolism; however this component

is normally lacking in mining wastes. Another key factor is the need for an anaerobic

environment. It is unlikely, therefore that both T. fenooxidans (an aerobe) and

Desulþvibrio desulphuricans (an anaerobe) will be active at the same time' The

environment will favour one or the other.

4.2 Macrobiological Effects

The possibilities and problems of aquatic disposal of mining wastes have received

considerable attention in Canada for a number of years. Indeed, in the mid-1970's

there was a flurry of publications dealing with environmental and biological

implications of such practices. Clark (1974) made a comprehensive review of the

effects of effluents from metal mines on aquatic ecosystems in Canada and provided

toxicity information on some 47 components of mining wastes on the biota of receiving

waters, including plankton, benthos and fish. Effects of specific toxicants such as

4-5


arsenic, cadmium and copper also were given special attention at that time (Penrose,

1974; Ray and Coffin, 1977; Black et al., 1976). Harvey (1976) noted in his report on

aquatic environmental quality in Canada, that despite the many problems with tailings

ponds, on-land disposal was preferred over direct dumping of mine wastes into the

aquatic environment. As an example, Harvey (1976) cited the long-term disposal of

tailings from one mine on L-ake Superior (Reserve Mining Co.) which introduced more

than five times the solids entering the lake naturally through shore erosion and rivers

input.

Macrobiological effects, separated for editorial convenience (but not functional

significance) from microbiological effects, may be placed into four major categories:

turbidity, sedimentation, toxicity, and contamination. Each is discussed in turn.

Impacts related to turbidity include the biological effects of reduced water transparency

on primary and secondary production as well as effects on respiration, feeding and

other behaviour of water column organisms in both "standing" (lakes, reservoirs, ponds)

and flowing waters. Sedimentation effects embrace phenomena associated with settling

and smothering of benthic organisms on lake and river bottoms. The discussion on

toxicity will inclucle a wicle range of lethal, sub-lethal and behavioural effects of trace

metals or acicl-generating materials on freshwater biota. Under the heading

contamination, the uptake and bioaccumulation of trace metals will be considered, as

well as their biomagnification; especially by inclusion of metals in short, direct food web

linkages.

Finally, because the effects of aquatic disposal of mining wastes can vary so greatly

depending on the type and conditions of receiving waters, it will be necessary to

comment briefly on some significant interactions in this area.

4,2.L Turbidity Effects

Many inlan<l waters of Cana<la were subjected to much higher levels of turbidity during

late phases of cleglaciation than occur today, so the biota have had a long evolutionary

history with turbidity in this area. Despite this, the possibilities for adaptation are not

great. Hence long-term changes in turbidity usually result in major changes in species

composition and abundance. Effects on primary producers are expressed mainly, but

not entirely, through reduction in light penetration by high turbidity; in lakes (Wetzel,

1983), and also in rivers such as the Fraser (Northcote et al., L975; Northcote and

4-6

LIMNOLOGICAL AND BIOLOGICAL CONSIDERAT¡ONS

I-arkin, 1983). In addition to particle concentrations, particle shape can have important

consequences; in particular the scouring action of suspended particles on periphyton

(attached algae) in streams and rivers can be greater if the particles are derived from

crushed rock in mining operations rather than from natural alluvial sediments.

Furthermore, the respiratory and filter-feeding structures of many planktonic as well as

benthic invertebrates are hampered by high suspended sediment levels in lakes and

streams.

The acute and chr<lnic effccts on fish of long-term exposure to stlspended sediment are

well clocumentecl (Wallen, 1951; Hebert and Merkins, 1961; Vinyard and O'Brien,

t976 Noggle, 1978 Sigler, 1981; Gardner, 1981; Crouse et al., 1981; Mcl-eay et al.,

1987; Servizi and Martens, 1987). Even relatively brief exposure (minutes to hours) to

suspended sediment levels over a few hundred mg.Lr can cause significant mortality in

some salmonid fishes, and longer exposure (several days) at levels in the range of a few

thousand mg.Lr can produce obvious gill damage and severe mortalities (often >507o)

in rainbow trout. Moderate increases in turbidity often cause an increase in ventilation

and orygen consumption rates in fish (Horkel and Pearson,1976) and though there may

be no severe mortality as a result, there are energy costs involved. Turbidity effects that

are negative to some species (e.g. salmonids), may be positive to others, such as

cyprinicls (Graclall anct Swenson, 1982). Some species such as the arctic grayling can

survivc wcll given high suspencled sediment levels, l)ut may sh<lw signs <lf stress in <lther

physiological ancl behavioural interactions (Mcl-eay et a1.,1983; 1984). Similar effects

have been reported for yearling coho salmon and steelhead trout (Redding et al., 1987).

Exposure to short-term (1 hour) pulses of suspended sediment can result in territorial

breakdown, reduced feeding ability, and increased gill-flaring ("cough" response) in

stream-dwelling young coho salmon (Berg and Northcote, 1985). Indirect effects of

suspended sediment on fish, especially salmonids, include reductions in growth rate and

delays in migration.

4,2.2 Sedimentation Effects

The blanketing of lake or stream bottoms by sediment, if continuous, inhibits

colonization and production of periphyton ancl woulcl likely have similar effects on

macrophytes. The f<¡o<l supply (bcnthic algae) to many forms of benthic invertebrates

may thus be reduced by smothering, and the habitat for epifaunal forms which utilize

spaces between rocks or other bottom materials may also be degraded. Infaunal groups

4-7

LIMNOLOGICAL AND BIOLOGICAL CONSIDERAT¡ONS

such ÍN aquatic earthworms may not be affected however. Stretches of sand or other

fine secliment depositecl in streams may impede or block upstream movement and

hence colonization of stream insects (Luedtke and Brusven, 1976).

Perhaps some of the most serious detriments of sedimentation on salmonid fishes are

effects on egg and alevin developing in stream or lake gravels, and on fry emerging

from these rearing habitats. Changes in survival are directly related to the amount of

fine particle size materials (< -1-3mm) which are included in the spawning gravel.

Relatively small changes in the amount of fine materials can have large effects on

survival. For example, levels of about SVo by volume in spawning gravel have little

effect on survival of pink salmon eggs, but levels of l}Vo can reduce it by up to 50Vo.

Emergence of salmonid late-alevins or fry from rearing gravels may be greatly inhibited

by sedimentation and mortality at that stage can be severe.

Sedimentation can also have in<lirect effects on fish, mainly by alteration and reduction

in cover which thereby increases their predation risk (see for example Alexander and

Hansen, 1983).

4,2.3 Toxicity Effects

As noted previously, there are nearly 50 components in mining wastes which can have

toxic effects on biota in receiving waters (Clark, 1974). Only some of the more

important and common of these are reviewed here with their acute and chronic effects

on survival as well as their sub-lethal, physiological and behavioural manifestations. To

facilitate later location of the information available, the components covered will follow

the sequence: acid mine waters and aluminum (by themselves or together), arsenic,

cadmium, chromium, copper, cyanicle, iron, lead, mercury' nickel, tin, zinc, and mixtures

(rtrganized under the bi<¡tic groups: phytoplankton, periphyton' macrophytes,

zooplankton, zoobenthos and fish). No attempt was made to cover effects on higher

vertebrates. I-aboratory, lake ancl stream (or river) results are included. For details on

many components, Chapter 3 of the Canadian Water Quality Guidelines (CCREM'

1987) should be consulted.

4.2.3.1 Plrytoplanldon

L¡rwer phytoplankton cell density, procluction, and species cliversity were recorded ín

two northern Ontario takes affected by acid mine wastes, compared to an unaffected

4-8


one (Johnson et al., 1970), but possible interactions with aluminum were not measured'

Aluminum can have direct toxic effects on phytoplankton in concentrations as low as

135 pg.Ll (Bohm-Tuchy, 1959), but in acidic waters may act through phosphorus

limitation as Nalewajko and Paul (1985) have shown in Ontario Precambrian Shield

lakes where significant ctecreases in photosynthesis were demonstrated at an aluminum

concentration of 50 ¡^lg'Ll.

Arsenic, in the presence of high nutrient (N, P) concentrations, did not inhibit high

algal biomass in the tube experiments of Brunskill et al. (1980). Within the ambient

concentrations used (< 160 tt|.ur), conway (1978) found no detrimental effects on

growth or on micronutrient utilization for the planktonic diatom Asteríonella formosa.

Furthermore, Planas and l¿marche (1983) found no effect of arsenic on phytoplankton

communities developed under various nutrient conditions. Interestingly, Baker et al'

(19g3) reported that arsenic methylation occurred in mixed green algal phytoplankton

cultures, suggesting an additional source for the formation and cycling of organo-arsenic

compounds in freshwater ecosystems. Vocke et al. (1930) noted arsenic toxicity at

levels ranging upwards from 48 pg.V' f.ot Scenedesmus obliquus.

Cadmium is rapidly sorbed by the common planktonic cliatoms Asterionella formosa and

Fragílaria crotonensis, actively by the former but passively by the latter (Conway and

Williams, tgTg). As concentrations were increased in the 2-9 Pg'Ul range' the growth

rate of the former decreased (Conway, 7978; Conway and Williams, 1'979) but was

unchanged in the latter.

Although copper has long been known to be toxic to phytoplankton at very low

concentrations, for example in the 10{0 to 10-12 M range (Stumm and Morgan, 1981),

there are suggestions that secretions of complexing ligands by some algae can

ameliorate such toxicity (Van den Berg et al., 1979). For the blue-green ph¡oplankter

Aphanizomenon flosaquae, Wurtsbaugh and Horne (l9SZ) have shown a linear

inhibition of N and C fixatíon as well as pigment accumulation between 10 and

30 t¡g Cu'L-r.

Zinc concentrations in the 10-20 ¡^rmol.Ll affect growth of. Chlamydomonc$ variabilis

with important interactions with phosphorus and pH (Bates et al., 1983, 1985;

Harrison et al., 1986).

4-9


Synergistic effects of metal mixtures on phytoplankton are now starting to be examined

(Wong et al., 1982).

4.2.3.2 Macrophytes

Aluminum concentrations of 2.5 mg.Lr brought about a 507o reduction in root growth

of Eurasian milfoil (Myriophyllum spícatum) at apH near neutrality (Stanley, L974).

The Northwest Miramichi River system of New Brunswick has been affected by copper-

zinc mining pollution since the early 1960's (Besch and Roberts-Pichette, 1970). After

an eight year period, mine water discharges seriously reduced or eliminated the riparian

vascular flora. Submerged macrophytes were the most sensitive group with the

horsetail Equisetum aruense the least sensitive species.

Revegetation of 32year old tailings in Mandy t-ake (Hamilton and Fraser, 1978), rich

in sulphides, took place by sedges (Carex spp.), riverweed (Podostemun ceratophyllum)

and spike rushes (Eteocharis spp.), whereas macrophyte species in areas away from the

tailings were more diverse; mainly cattails (Typha latþlia), yellow pond lily (Nuphar

varíegatum), water smartweed (Potygonum amphibium), bullrushes (Scfipru spp.) and

pondweed s (Potamogefon spp.).

4.2.3.3 Zooplanldon

Toxic effects of aluminum, in combination with acid waters, often appear at

concentrations below 1 mg.Ll (Biesinger and Christensen, 1972; Shephard, 1983;

Havas and Likens, 1985; A¡ts and Sprules, 1987; see also CCREM, 1987). Although

estuarine zooplankton are resistant to relatively high arsenate concentrations (up to

100 pg. Lt), indirect effects may be much more severe (Sanders, 1986).

Cadmium appears to be especially toxic to zooplankton with levels as low as 10{5 M

showing effects on Daphniø (Stumm and Morgan, 1981). Population decline and

reproductive depression have been reported in the 0.2 to 4 pg'L;r range (Marshall,

L978; Marshall and Mellinger, 1980; Marshall et al., 19S1) but the effects were

decreased at low pH (Lawrence and Holoka, 1987).

During molt, Daplmia pulex has a significantly higher mortality when exposed to

chromium at 0.56 mg.Lr (tæe and Buikema, 1979).

4-10


Toxicity of copper for Daphnía magnøput the adult 3-week LCro at 0'044 mg'Ll and a

50To reproductive impairment at 0.035 mg.Ll (Biesinger and christensen, 1972), but

both are increased to 0.26mg.L-1 when zmg-Ll of the chelator NTA is added

(Biesinger et al., Lg74). Winner and Farrel (1976) give chronic toxicity values of copper

for four species of DaPhnia.

Synergism of toxicity in multimetal mixtures has been known for some time in

freshwater organisms (Anderson and weber, 1975) and has been clearly demonstrated

in freshwater copepocls, primarily Cyclops spp' (Borgmann' 19t10)'

4.2.3.4 Zoobenthos

Significant (37Vo)mortality of chironomid larvae (Tanytarsus dissimílis, second and third

instars) occurred at aluminum concentrations of 0.8 mg !r at pH 6.8 (I-amb and Bailey,

1981). Effects of cadmium were reported for lakes (Andersson and Borg, 1988) and

streams (Stephenson and Mackie, 198s). Toxic effects of copper in relation to pH have

been examined for two species of amphipods by de March (1979,1983)' Earlier studies

on copper effects include those of Arthur and Læonard (1970) and Peterson (1978)'

clements et al. (1ggg) have recently shown that experimental stream results of copper

and zinc toxicity compare well with field results.

I-æacl concentrations as low as 19 t g.vt providecl significant increases in snail

(Lymnaea patustrís) mortality but not growth rate (Borgmann et al', 1978)' Mercury

exposure reduced artificial stream biomass and diversity (Sigmon et al., 1977). Other

effects of heavy metals on stream survival and community structure of insects are given

by Hal et al. (1988), Winner et al. (1980) and Burton and Allan (1986); see also Havas

and Hutchinson (1932) and waterhouse and Farrell (1985). Acute toxicities to

amphipods of binary mixtures of metals are considered by de March (1988)'

4.2.3.5 Físh

More general aspects of the toxic effects of mine wastes on freshwater fish have been

reviewed from the eastern to western regions of Canada (Elson et al., 1973; Elson,

1974; Somers an{ Harvey, 1984; Alderdice and Mc[.ean, 1982). A variety of

physiological and behavioural effects have been suggested, including altered

chemoreception (Hara, Lg72, lgSt: Brown et al., L982), spermatogenesis (Cochran,

1987), bone development (Hamilton and Reash, 1988), fecundity (Reash and Berra,

4-11

LIMNOLOGICAL AND BIOLOGICAL CONS¡DERATIONS

1986), temperature selection (Peterson,l976), vertical positioning (Scherer, 1976), and

susceptibility to bacterial infection (Pippy and Hare, 1969). Sensitive indicators of

metal pollution using fish have been suggested (Roch et al., 1982) as have the

interactions between genetic and environmental factors in the development of

resistance to such pollution by fish (Swarts et al., 1978).

Explanations for the decline or disappearance of some species of salmonids from

headwater lakes and streams now include complex interactions between acidity, low

calcium and toxic aluminum concentrations, as illustrated for brook trout (Cleveland et

al., L986; Siddens et al., 1986; Mount et al., 1988a" b). Changes in gill structure also are

involved (Tietge et al., 1988). Similar or complimentary investigations are being

conducted on other salmonids such as rainbow trout (Neville, 1985) and lake trout

(Gun and Noakes, 1987).

Early work on toxicity of arsenic to fish was conducted by Alderdice and Brett (1957).

Acute toxicity of arsenic for several freshwater fishes is in the 10-15mg'Lr range

(CCREM, 1987) but chronic toxicities are reported at much lower levels.

Acute toxicity levels for cadmium are in the 7-3 ttg. Ll range for rainbow trout, with

slightly lower chronic toxicity values. Both are inversely dependent on water hardness

(CCREM, 19S7). Possibly because rainbow trout are the piscine equivalent of the

laboratory tat, alarge body of detailed physiological work has been done on this species

with respect to cadmium (and other heavy metal) toxicity (see for example Chapman,

1978a;Roch and Maly, 1979;Pärt and Svanberg, 1981; Majewski and Giles, 1981; Giles

1984; I¡we-Jinde and Ntimi, 1986; Reid and McDonald, 1988; Giles, 1988).

Nevertheless there have been numerous studies on cadmium toxicity in other salmonids

such as brook trout (Benoit et a1.,7976; Sangalang and Freeman,1979; Hamilton et al.,

1987a,b), Atlantic salmon (Peterson et al., 1983; 1985), and chinook salmon (Finlayson

and Verru e, 1982). Effects on catostomids, i.e. suckers, (Duncan and Klaverkamp,

1983; Borgmann and Ralph, 1986) and cyprinids have not been overlooked (Pickering

and Galt, 1972; Sullivan et al., 1978; McCarty et al., 1978; Houston and Keen, 1984;

Andros and Garton, 1980).

In addition to the few references on chromium toxicity reported by CCREM (1987)' the

study by Adelman et al. (1976) on two species of cyprinids should be noted.

4-12


Acute and chronic toxicity of copper seem to be reasonably well established for rainbow

trout (Harrison, 1975;Giles and Klaverkamp, 1981; CCREM, 1987) and other studies

on this species have consídered effects on olfaction (Hara et al., L976), bioenergetics

(t ett et al., Lg76), hematocrit (Waiwood, 1980), acclimation (Dixon and Sprague, 1981;

I-aurén and McDonald, 1987a, b), avoidance (Giattina et al., 1982), and growth and

survival (Seim et al., 1984). Nearly as well studied are lethal and sublethal effects of

copper on Atlantic salmon (Sprague, L964a; Sprague and Ramsay,1965; Sprague, 7965;

Sprague et al., 1965; Zitko et al., 1973) as well as its effects in this species on avoidance

(Sprague, 1964b) and on migration (Saunders and Sprague, 1967; Sutterlin and Gray,

lg73). Similar studies on other salmonids include those on brook trout (McKim et al.,

t970; McKim and Benoit, 1g7l; Drummond et al., 1973), coho salmon (Waldichuk,

1976) and on sockeye salmon (Davis and Shand, 1978). Copper effects on non-

salmonids have been examined in striped bass (Bohammer, 1985)' white suckers

(Munkittrick and Dixon, 1988), cyprinids (Mount and Stephan, t969; Kleerkoper et al.,

lg71,1973;Tsai and McKee, 19S0) and in an ictalurid, i.e. catfish (Brungs et al., 1973).

For cyanide, acute and chronic toxicities as well as several other effects and interactions

are covered in the literature cited by CCREM (1987). The studies by Broderius et al.

(1g77),Kimball et al. (1978), Iæduc (1978), Kovacs and læduc (1982a,b) provide usetul

examples.

In addition to the few studies on iron toxicity reported by CCREM (1987)' those of

Brenner et al. (1976) on the common shiner and of Smith and Sykora (1976) on brook

trout and coho salmon should be consulted.

Long-term effects of lead exposure have been described for brook trout (Holcombe et

al., \976; see also Dorfman and Whi¡vorth, 1969) as wetl as its chronic and sublethal

effects on rainbow trout (Hodson et al., 1980; Sippel et al., 1983; see also Hodson et al.,

lgg2). Blood characteristics of several fishes have been used as an indicator of harmful

exposures to lead (Hodson, 1976a;Hodson et al., 1977,1978: Schmitt et al., 1984)'

Acute toxicity levels for forms of mercury have been reported for rainbow trout fry and

fingerlings (Wobeser, L975; see also CCREM, 19S7). Mercury is said to block taste

responses in Atlantic salmon parr (Sutterlin and Sutterlin, 1970).

Many of the studies on copper toxicity reported previously, especially those on brook

trout and Atlantic salmon, also provided information on zinc toxicity. Additional work

4-13


oii zinc includes that by Sprague (1963) and Hodson (1976b) for rainbow trout, effects

on Atlantic salmon (Hodson and Spragu e, 1975;Zitko and Carson, 1977) and on brook

trout (Holcombe et al., lg7g) and sockeye salmon (Chapman, 1978b; Boyce and

yamada" lg77). Zinc is also known to affect the immunity of fish to viral and bacterial

infection (Sarot an¿ Perlmutter, 1976) as well as their growth, sexual maturity and

reproduction (Pierson, 1981).

Efforts to demonstrate genetic selection for zinc tolerance in laboratory (flagfish) and

witd (common shiner) populations of fish were not successful (Rahel, 1981).

4.2.4 Contamination Effects

There is an enormous volume of literature on the uptake, bioaccumulation and

biomagnification of many metallic components originating from mine wastes discharged

into freshwaters, as well as from natural inflows. The study by Wagemann et al' (1978)

provides an example for arsenic, giving concentrations in the sediment, water and biota

compartments. Allen (1936) reviews the bioavailability and bioaccumulation of toxic

metals in the biota of several large Canadian lakes and rivers. Examples of lead'

mercury, nickel and tin accumulation in phytoplankton are given, respectively' by

Denny and Welsh (Lglg), Rudd and Turner (1983), Watras et al. (1985) and Wong et

al. (198a).

Metal contamination in aquatic macrophytes has been studied by Franzin and

McFarlane (1980), Marshall et al. (1983), Campbell et al. (1985), and Andersson and

Borg (1988).

Accumulation levels of several metals in zooplankton are reported by Denny and Welsh

(Lg7g),watras et al. (1985), Bodaly et al" (1987) and Jackson (1988).

For the benthic community, bioaccumulation data for several trace metals may be

found in the publications of Smith et al. (1975), Bindra and Hall (MS 1973)' Rudd et al'

(1980), Tessier et a!. (19t14), Evans anrl l¿senby (1983), Van Duyn-Henclerson and

Lasenby (1986), Jackson (1983) and Evans et al. (1988)'

For fish, correlations between sediment concentration of several metals and that in fish

from several Ontario lakes have been reported (Johnson, 1987)' Cadmium uptake by

fish has been observed in several systems (Atchison et al., L977; Ramamoorthy and

4-14


Blumhagen, 1984; Andersson and Borg, 1988). Accumulation of chromium by fish

(Buhler et al, lg77) and of copper (Duthíe and Carter, 1970) has been recorded' One

of the most intensively studied trace elements in respect to uptake by fish has been

mercury (Gillespie and Scott, 1972;Uthe et al., 1973; Scott, 1974; Reinert et al., 7974;

I-aarman et al., L976; McKim et al., 1976; Scherer et al., 1976; Hartman, 1978;

Huckabee et al., L971;Phillips and Buhler, 1978; McFarlane and Franzin, 1980; Rudd

et al., 1980; Rodgers and Beamish, 1981, 1983; MacCrimmon et al', 1983; Walczak et

al., 1986; Hecþ et al., 1987;Bodaly et al., 1987).

zinc uptake by fish has been studied by Hodson (1975), Atchison et al. (L977),

Ramamoorthy and Blumhagen (1984), and Spry et al. (1988)'

4.2.5 Receiving lVater Conditions

The physical, chemical and biological condition of waters proposed or being used for

subaqueous disposal of mining wastes will have major implications on the suitability of

such systems for that purpose. Perhaps the first major consideration is whether or not

the waters are "standing" or moving, i.e. is the proposed site a lake, pond, reservoir or is

it a stream or river. Responses of these two major categories to subaqueous disposal

can be profoundly different.

The second major consideration is the question of scale; largely revolving around

morphometric features of the receiving system. The size of the receiving water in

relation to the quantity of input is critical. In the case of lakes or reservoirs, the surface

area to volume relationships are important but so are other morphometric features such

as fetch, shoreline development, average basin slope, and ma¡rimum depth' In a small

lake, a high tailings input and sedimentation rate can have severe impacts on biota and

recreàtional use (Osborne, Lg76). For rivers, another set of morphometric parameters

must be used including, for example, width, cross-sectional area, bank slope and

maximum depth. The basin (lakes) or channel (rivers) morphometry will, in part, set

the discharge characteristics, which wilt have major bearing on site suitability.

Short-term, seasonal as well as annual variation in retention time (or its converse;

flushing rate) must be considered for lakes and reservoirs, as must discharge

hydrographs for rivers.

In lakes and reservoirs the existence, extent and timing of seasonal stratification and

mixing patterns must be known as these features will profoundly influence disposal

4-15


considerations. It must be known therefore whether or not the lake is holomictic or

meromictic, dimictic or polymictic, and the depths, temperatures and other related

features of the stratifications must also be known. For rivers, the mixing patterns are

different but there may well be marked cross-sectional and "side" differences to be

considered. Because rivers flow, does not necessarily mean that mixing is instantaneous

or spatially localized. Pollutant inflows into large rivers often impart a "one-sidedness"

in characteristics for kilometres downstream.

There are also many chemical parameters to consider. The non-mixing

(monimolimnetic) waters of meromictic lakes often are markedly different in chemical

characteristics compared to the overlying mixed (mixolimnetic) waters. There can be

major differences, therefore, in clissolved gases, minerals, alkalinity, conductivity and

salinity, between upper and lower layers. These differences can have major influence

on the turbidity, seclimentation, toxicity and contamination effects considered

previously. Even in holomictic lakes there can be great seasonal differences in the

extent, depth and effectiveness of mixing and thereby in the chemical conditions noted

above.

Biotic interactions, whether at depth in the sediments, at the mud-water interface, in

the water column itsell or at the water-air interface, can be of great importance in

effecting metal toxicity and contamination. The benthic infauna can effect major

changes in mineral sedimentation, stratification and availability within bottom deposits'

Some benthic invertebrates such as mysids (a type of freshwater shrimp) and

chaoborids spend daylight hours as benthíc feeders and rise each night up to near-

surface waters. Through feeding, vertical migration and preclatory interactions, they

can act as concentrators and transporters of materials between bottom and water

cr¡lumn compartments of lakes. For example, certain heavy metals (zinc, lead and

cadmium) are transported via the vertical migration of Kootenay L,ake mysids, which

are consumed by kokanee salmon (Evans and I-asenby, 1983; Van Duyn-Henderson

and l-asenby, 1986). A vertical migration of 30-40 m is not uncommon. Such metal

uptake by fish suggests only a very short food chain linkage to man. It seems clear that

the community structure and food web structure of lake and river ecosystems also may

play major roles in their functional processes, concentration or cycling of elements and

thus in problems of toxicity and contamination.

4-16

5.0 SELECTED CASE STUDIES

5.1 lntroduction

A number of case studies which either have or are currently utilizing freshwater

subaqueous disposal are outlined in Tables 5-1 and 5-2 and described below. The sites

chosen represent a range of ore types and receiving environments. Not all mines

utilizing subaqueous disposal practices have data available which are adequate for

understanding the impacts of tailings input, although considerable monitoring data may

exist. We have endeavored to include as much detail as possible on the sites listed in

Tables 5-L and 5-2; however, as will become evident, in many cases only limited

information is available. Lake disposal at Brucejack, B.C. (Newhawk Gold Mine's

Sulphurets Property) is being considered for startup in 1989. Bearskin l-ake at the

Golden Bear project (Noramco/Chewon) was initially considered for lake disposal but

was subsequentþ rejected. Some information on these trvo projects is also included

below.

5.1.1 British Columbia Mine Sites

5.1.1.1 Babine Løke (Granßle Mines)

Noranda Minerals Inc. acquired the Granisle Mine and its operating and ancillary

surface facilities fromZapata-Granþ Mining in the late 1970's. The mine is currently

operating along with Belt Copper, and together the two mines form Noranda's Babine

Division. Granisle is a copper-mineralized po¡phyry deposit comprised of a central

bornite-chalcopyrite zone grading outwardly to chalcopyrite; a pyrite halo surrounds the

copper-rich zone.

The Granisle Copper mine-mill operation is on McDonald Island in Babine l-ake, with

tailings ponds occupying the area between McDonald Island and Sterrett Island to the

south. The mine-milling complex has a closed tailings system with discharge to the

sectioned-off area of the lake and water recycled from the tailings pond for reuse in the

mill. The lake is monitorecl in the area just off the No. 2 dam fronting east on the lake

(between McDonald and Sterrett Island) and the settling dam, fronting west. A lake

station, about 1 km offshore is monitored as a control.

5-1

SELECTE

Table 5-1

Selected Case Studies, lake Disposal of Mine Wastes

D CASE STUDIES

lake/Mining ComPanY or ProPeñY Ore Type

British Columbia

Canada

lnternational

Babine l-ake. Granisle Mines (Noranda lnc.)Bearskin Lake-*:-'ñ;ã;;o/chevron Mtnerats Ltd. (Gotden Bear operatlng co.)

Benson Lake. Cominco's Coast CoPPer MineBruceiack Lake. 'Newhawk Gold Mines Ltd'Buttle Lake-- .- Westmin Resources (Western Mlnes) Ltd'

Kootenav Lake'*:-'öót i"cã's Bluebell Mine at Riondel (east ban[). Dragoon Resources at Ainsworth (west banx¡'

Pinchil-ake. ComincoSummit Lake. Scottie Gold Mines Ltd.St. Mary's River/KootenaY River

. Sullivan Mine - Cominco Ltd

Cu

Au

Cu

Au

Cu/Pb/Zn

Pb/7nPb/znlAg

Hg

Au

Cu/Ni in massive sulPhides

PblZn

Cu/Ag/Au

Cu/ZnlPb

Fe Taconite(siliceous)

Fox Lake, Manitoba. Farlev & Sherridon MineGarrow l-atä, Uittte Cornwallis lsland, N'W'T

. Polaris Mine/Cominco Ltd'Mandv Lake, Manitoba- .-'MandY Mine/Hudson BaY M & SAnderson Lake, Manitoba

. HudsonBaYM&S

Silver Bav fl¿ke Superior, Minnesota, U'S'A'). Resèrve Minihg Co.

1. FormerlY Davld Mlneral¡.

5-2

SELECTED CASE STUDIES

Table 5-2

Selected Case Studies, Flooded Pits and Shafts

rNote: Prolects ln conceptual 3tagog.

Water quality monitoring data for 1976 are shown in Table 5-3. Analyses for samples

taken off the face of the two dams correspond closely to those for the control station in

the lake. I¿ke water was only moderately alkaline, with relatively low dissolved solids

and low suspended solids. All metal values were low, as was sulphate, and cyanide was

not detected. Overall data indicate water of high quality. The pH range for samples

taken adjacent to the dams was above similar values for the control station; however

values were still within the biological range. Increases in copper levels (over

background) were evident in samples taken at the dam faces, but again values were still

low.

The Bell Copper tailings pond is in complete recycle with mill operations and ditches

collect seepage from the various dams for return to the pond. The company monitors

water quality parameters at three locations in Babine Lake, immediately offshore of the

two major tailings dams, fronting in the lake and in Rum Bay, and a third location some

distance north, offshore from the concentrator. Surface, bottom and mid-depth waters

are monitored.

Typical analyses are shown in Table 5-4. Anal¡ical values indicate a moderately

alkaline lake, of high clarity, with minimal dissolved solids and particularly low sulphate

values. Monitoring for metals, both total and dissolved, has indicated low levels below

the <letection limits of the analytical methocls. Arsenic and cyanide levels also were

low. In fact, analytical values for individual parameters were virtually identical for all

Ore TypeMinlng Company

Endako Mines Divislon, Placer'Dome lnc. (flooded pit). Fraser Lake

Equity Silver Mines Ltd. (flooded Southern Tall plt). Houston

Phoenix Mlne (decommlssloned open plt). Greenwood

Cinola Gold, City Resources (Canada) Llmited*. Queen Charlotte lslands

Mo

Ag

Cu

Au

5-3


Tablo 5-3

Water Quality Gharacterlstlcs of Bablne Lakein the Environs of Granlsle Copper Operations

East of No.2 Dam West of Settling DamWNWof Settling Dam

(3/4 mile)

Avg. Range Avg. Range Avg. Range

Totalsolids

Suspended Solids

pH

Copper

Zinc

lron

Totalcyanide

Sulphate

66-170

1-36

7.0-8.5

0.00s-0.050

0.001-0.025

0.005{.050

5.8-6.2

96

I7.5

0.023

0.016

0.010

<0.2

6

58-120

1-15

7.1€.3

0.005-0.030

0.001-0.015

0.005-0.01f

3.3€

83

4

7.7

0.014

0.008

0.006

<0.02

4

58-140

1-1 1

7.2-7.6

0.004-0.0'14

0.001-0.010

0.005-0.012

77

3

7.4

0.008

0.006

0.007

<0.02

6 3-8.3

Note: Unltg lor att values shown are ln mg/L except pH.

5-4


Table 5-4

water Quality characteristics of Babine Lake in thelmmeOiãté eñv¡ro¡ís of the Bell Copper Mine-Milling Operation

rNote: Value lndlcated ropros€nts result¡ lor both total and dlg¡olved motals.

three sites and for each of the three depths at each site. Overall results do not indicate

any impact on lake water qualitY.

5,1.1.2 Bearskin Lake, Noramco/Chevron Minerals Ltd'

North American Metals Corp. and Chevron Minerals Ltd. are joint venture partners of

the Ggldcn Bear gotd project near Telegraph Creck. It is schecluled for prodtrction

starr-up in the thir¿ quarter of 1989 ancl will use combined open pit and underground

mining. The 360 tonnes per day mill will use dry grínding, fluidized bed roasting and

carbon-in-pulp cyanide leaching for gold recovery.

The original concept was for lake clisposal of tailings, via a barge, into Bearskin [-ake,

although perhaps not year round (Stage I Environmental Impact Assessment, Vols. 1,, 3,

July 1987). Moclelling and computer simulations indicated that Hg and Pb would

Rum Bay (No. S Dam)West of ConcontratorNo.1 Dam

Location

Middle Bottom Surlace Middle BottomSurface Middle Botlom Surlace

pH

Dissolved solids (mg/L)

Suspended solids (mg/L)

Turbidity (JTU)

Oiland grease (mg/L)

Sulfate (mg/L)

Cu* (mg/L)

Zn* (mg/L)

Pb* (mg/L)

Cd* (mg/L)

Ni* (mg/L)

As* (mg/L)

Ag* (mg/L)

CN, (mg/L)

7.7

60

1

4.0

2.4

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.7

68

1

3.9

2.1

2

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.6

47

1

3.7

1.8

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.7

70

1

4.0

2.7

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.7

66

3

4.0

1.8

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.7

70

f3.7

3.2

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.7

55

<1

4.'l

2.6

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.6

65

<1

3.7

3,2

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

7.6

62

<1

3.5

2.2

3

<0.01

<0.01

<0.01

<0.01

<0.01

<0.02

<0.01

<0.01

5-5


periodically exceed water quality guidelines. When Hg and Pb metal levels exceeded

allowable limits, the tailings would be sent to on-land disposal sites, which would also

be used during early operation.

The objectives for the maximum concentration of metals in the lake are:

Hg - 0.1 þElL

Pb - 5 pelL(publishecl provincial objective set in other watersheds)

2 pe/L (federal guideline)

At the time of this report, the tailings disposal concept had changed to eliminate any

lake disposal of tailings.

5.1"1.3 Benson Lake (Coast Copper Co./Comínco)

Cominco's Coast Copper mine, located near Port Hardy, 8.C., exploited two

underground deposits - Benson l¿ke and Coast Copper. The mine operated from

August 1962 to January 1973 when it closed down because of unfavourable economics.

The tonnage mined over the operating interval totalled 3.6 million tons.

Although rated at 750 tpd, the mill often processed 850 tpd. Copper concentrate was

produced from high grade ore (mean grade 2.02%o Cu). In March L963, a magnetite

recovery plant began operation to produce iron concentrate (64'65Vo Fe) from iron

plant feed assaying}g%oFe. The iron plant shut down in September 1970 because

sulphur content in the concentrates exceeded specifications.

The mill discharged tailings under permit into Benson l-ake. Terrain constraints

precluded land disposal into a conventional impoundment. Figure 5-1 shows Benson

I^ake bathymetry, Secchi disc stations and the initial tailings discharge location. A

typical mineralogical composition of the tailings is listed in Table 5-5. From Table 5-5,

the sulphur content of the tailings is estimated to be approximately lVo.

Accorcling to reports hy DOE/Fisheries Service ancl Environment CanadalEPS-PR,

excessive turbiclity was an immediate and lingering problem throughout the mine's

operation (Benson t-ake Monitoring Data (pH, Clarity/Secchi, turbidity), 1961-68).

The tailings contained a slow-settling colloidal fraction that the Cominco and

government reports did not identify.

5-6

lRoq¡ng R¡ver

BensonRiver

I.AKE-¡o

/oo Discharge Location

9o

Upper Benson Rlver Approx Scole l"=llOOfl( All reodings ín feel )

M orch 16, l970Cominco Secchi disc stolions

Flgure S-t Benson l¡ke Showlng Contours, SecchlDlsc Statlons, and Dlsctnrge Locatlon (Kussat et al" 1972) (@


Table 5-5

Estimated Minerats Composltion for Goast Copper Mine Tailingsl

MineralEstlmated Percentage

ln the Tailings

32.0027.003.004.000.500.500.500.300.30

28.80.11

.101.001.00

.89

Epidote

Feldspar.DiopsideActinoliteChlorlte..

Sericite..Magnetite..ChalcopyriteBornite......Pyrrhotite".Pyrite..........Unidentified

11960 Comlnco lntornrl momo.

Cominco attempted unsr¡ccessfully to mitigate the problem through the use of

flocculants. At the request of the Fisheries Service, Cominco twice moved the tailings

outfall to deeper areas of the lake. A summary of these efforts is listed belowl:

11967 Comlnco lntern¡l memo.

Despite these measures, the turbidity problem remained, particularly during the winter

months when the lake was isothermal" Extensive limnologic surveys were carried out in

1970 and 1971, in an attempt to understand the interplay between temperature

TailingsOutfallLocation Dates

Ore TreatedShort Tons

LakeDepth m

DownpipeLength m

lnitial Aug./62 - Sept./al

Sept./6zl- Nov./70

550,000

1,800,000

42.7 30.5

First move;305m downlake plus raft

45 30.5

Second move; a further305m down lake

Nov./70 - Jan./73(mine closed)

900,000 48.8 45

5-B


stratification and turbidity in the water column. Temperature profiles collected at

Stations 7, 2, and 3 in July 1970 (Figures 5-2, 5-3 and 5-4) indicated typical summer

stratification with the thermocline occurring between about 7.5 and 20 m depth.

Subsequent surveys outlined isothermal conditions on March 16, 197I and again on

September 9,7971, indicating in the latter case an early fall turnover. On these dates,

suspended tailings solids were found to be present throughout water column to a depth

of 50 m at Station 2, andthis finding was confirmed at two additional stations. An EPS

report (Haltam et al., 1974) suggested that the seasonal thermocline acted as a density

barrier and prevented the colloiclal tailings in the hypolimnion from entering the upper

epilimnion. Subsequent homogenization of the lake water during the fall turnover

promoted dispersion of colloidal tailings throughout the water column (Kussat et al.,

lg72). Turbidity in the lake during the winter was also influenced by the input of

sediments from inflowing streams. There are no turbidity data prior to commencement

of discharg e in 1962 although Secchi disc readings were taken during the preceding

winter.

Few physical or water quality data other than Secchi disc, temperature and pH

measurements were collectecl during the limnologic surveys, and baseline studies prior

to the mine's start-up do not exist since none were required at the time of the mine's

1962 inception. However, some dissolved metals data are available. Measurements

reported by Kussat et al. (1972) suggested that the zinc level in the soft waters of

Benson lake was high (Table 5-6).

Table 5-6

Mean Concentrat¡ons of Some Heavy Metals in Jroutand Water from Benson and Maynard Lakes'

Specimen M€talsBenson L¡ke (ppm)

WalorBenson Lake (ppm)

Fish

Maynard Lake (ppm)Fish

Salmo clarki(cutthroat trout)

HgCuZnPbfr

<0.00005<0.005

0.06<0.01

0.1

0.26.50.1

0.1

0.1

0.38.50.1

0.1

t F¡gh ramplea taken March 16, 1971; water campte taken Febru¡ry 2, 1971; heavy meta! analyses by Comlnco'r Tralllaboratory. Data from Kussat ot al., 1972.

5-9

Temperoture oC

2 t4 t8

y 6, l97Oro

E

o'oo

r5 Morch 16,l97l

2

35 SeptemÞer 9, l97l

Figure 5-2 Benson Lake Thennal Stratiflcation, Station t (Kussat et al., 1972)

Temperoture oC

2 6 r8

57, t970

roMorch l6, l97l

?15E

;zoÔ'a,o25

30

35

SeptemÞer 9, l97l

Fìgure 5-3 Benson l-ake Thennal Stratiñcation, Station 2 (Kussat et al., 1972)

Temperolure oC

t4

5

ro July 6, l97O

€ r5

f20ooo25

30

35

Morch 16, l97l

SeptemÞer 9, l97l

Benson Lake Thermal Stratiflcatlon, Slatlon 3 (@Figure 5-4 (Kussat et al., 1972)


Biological investigations were initiated in February 1967 in an attempt to document

benthic fauna and determine the substrate composition in the deeper waters of Benson

[¿ke. At that time, tailings were found to cover the entire lake bottom and there were

no benthic organisms. Turbid waters were observed in the lower Benson River together

with a shallow deposit of tailings fines and coexisting benthic organisms.

No significant clifference was observed in heavy metal content between cutthroat trout

(Salmo clarkí) samples collected in 1971. from Benson and Maynarcl l-akes (Table 5-6),

despite the fact that the latter, a dammed lake immediately upstream of Benson l¿ke,

was not influenced by mining activity (although its watershed has been logged

extensively). The small sample size, however, (two trout per lake) rendered the

comparative results inconclusive. The coincidence of the high zinc levels in both fish

tissue and the soft water in Benson Lake, coupled with recognition that soft waters

enhance metal toxicity to aquatic organisms and previous observations of high zinc

levels in Benson River water, led to a recommendation for increased monitoring. In

November lgT3,after the fall turnover and some ten months after closure of the mine,

EpS conducted a survey of Benson l-ake to determine physical and chemical

characteristics, and to assess impacts to the biota (Hallam et al./EPS, L974). A marked

re¿ucti<ln in turbiclity was notecl, although total metal c<lncentrati<lns in lake water and

fish tissue were unchanged between 1971 an<t 1973. No significant bioaccumulation in

the food chain of heavy metals contained in the tailings appeared to be occurring.

5.1.1.4 Brucejack Lake (Newhawk Gold Mínes Ltd')

The Sulphurets property is a joint venture between Newhawk Gold Mines Ltd. (NPL)

and Granduc Mines Ltd., both of Vancouver. The property is located 56 km northwest

of Stewart in northwestern B.C.

Brucejack I-ake is located in the southeastern corner of the claims area at an elevation

of 1376 m. The lake is on a high plateau above the timberline and is ice-bound most of

the year. The total surface area of the lake is 81 ha with a length of 1300 m and an

average width of 623 m. The lake is 88 m deep at its deepest point, located close to the

centre of the lake. The total storage volume of the lake is approximately 29 million m3.

Underground mining is plannect at a production rate of 318 tonnes /day. Precious metal

recovery will be accomplished utilizing gravity concentration followed by flotation of a

silver-rich sulphide concentrate.

5-11


As tailings and waste rock were found to be potentially acid generating and because of

a limited amount of suitable land, tailings and waste rock will be disposed of

underwater in Brucejack Lake. The tailings, together with mine water, will bc

deposited using a submerged outfall located at depth (65 m) in the eastern basin of

Brucejack L¿ke.

Tailings solids are expected to settle on the lake bottom at a ma,rimum slope of 2Vo in

the deepest areas of the lake. Total tailings production from existing reserves is

expecrecl to be 550,000 t with 50Vo of the tailings solids being used for backfill in the

mine. The first site to be used consists of a small embayment adjacent to Brucejack

Creek which will be flooded following the installation of a hydroelectric intake dam on

Brucejack Creek. The second site selected consists of a small bay located in the

northwest corner of Brucejack l-ake. Dumping in this area will be restricted to a

5 month maximum open water period. An average 97 m3/hr combined tailings fines

and mine water will be gravity fed at approximately 6.9Vo solids to the submerged

tailings outfall.

Results from extensive laboratory testing of tailings supernatant and solids components,

including particle size distribution, metal analyses of tailings fractions and settling

velocities were used as input parameters to a one-dimensional finite difference modei

of Brucejack l¿ke for purposes of predicting the effects of deep lake tailings disposal.

The model used available estimates of minimum lake inflows, non-settleable

particulates and associated metal levels, together with conservative estimates of

relevant dispersive processes.

The suspended sediment model was run for a simulation period of 8 years to encompass

the estimate{ life of the mine, and predicted a generally increasing sediment

concentration leaving Brucejack L¿ke through the first five years or so of operating the

tailings disposal system. After this, the average annual concentrations levelled off

suggesting an equilibrium had been reached.

On an annual basis, the modelled sediment concentrations exhibited two distinct

phases. The first occurred at the autumn turnover when the sediment concentrations

increased markedly. The concentrations then remained approximately constant over

the winter and spring during the period of low inflow. The second phase occurred when

high summer runoff diluted the outflow concentrations to near background levels. The

outflow concentrations then increased marginally up to the autumn turnover event.

5-12


During the final year of simulation the peak suspended sediment concentration was

modelled at 13 mg/L, or roughly half the MMLER guidelines value of.25 mglL. The

particulate concentrations of eight metals (As, Cu, Fe, Pb, Hg, Mo, Ag and Zn) were

also predicted using the model. Table 5-7 lists the peak concentrations and flow-

weighted average annual concentrations leaving the lake during the final year of

simulation

The behaviour of the particulate metal concentrations followed similar time

distribution curves to the suspendecl sediment concentrations. The difference in

outflow concentrations between the summer and winter periods is dependent on the

relative metal concentrations of the creek inflows and the tailings discharge.

The modelling of lake tailings disposal demonstrated that outflows from Brucejack

[,ake will meet both fe<leral and provincial effluent quality criteria on a consistent basis.

On the basis of combined worst case scenarios for discharge quality and available

dilution in the receiving environment, only total particulate iron, lead, silver and copper

would be above the Canadian Council of Resource and Environment Ministers

(CCREM) recommended guidelines for protection of freshwater aquatic life inSulphurets Creek.

5.1.1.5 Buttle Lake (Westmin Resources Ltd.)

The discharge of mill tailings into Buttle l¿ke from Westmin's Myra Falls mine in

B.C.'s Strathcona Provincial Park represents a recent mining operation which has used

subaqueous lacustrine tailings disposal. The lake is the domestic water source for the

town of Campbell River, which is roughly 93 km from the mine. Pedersen and l-osher

(1988) reported on prior studies of chemical behaviour of Buttle L¿ke tailings; their

review has been supplemented below by recent information published by the B.C.

Ministry of Environment and Parks (Deniseger et al., 1988).

Buttle Lake is a large (30 km long by 1.5 km wide) lake which occupies a U-shaped

valley in an area of high relief on Vancouver Island. The south basin of the lake

reaches a maximum depth of 87 m, and during 1967 to 1984 received tailings via a

slightly submerged outfall. A relatively shallow sill 5 km north of the discharge site

effectively limits physical dispersion of the deposited material. By 1984 when tailings

disposal ceased, nearly 5.5 million tons of mill tailings had been deposited into the lake.

Tailings are currently being disposed on land.

5-13

Talble 5-7

Slmutated Metal Concentratlons at Outlet of Brucefack l¡ke- ÀttCr I years ot Operatlng the Tailings Disposal System

MMLER

UtglLl

500 - 1000

300 - 600

200 - 400

500 - 1000

Provlncial

Obfectives (ttgltl

100 - 1000

50- 300

300 - 1000

50- 200

0-5500 - 5000

50 - 500

200 - 1000

Flow-weighted AverageAnnualCìonc. (ttglL)

Total

5.3

4.3

590

9.6

0.094

f.5

1.5

11

Dissolved

2.5

1.1

35

0.61

0.084

'1.4

0.12

5.6

Particulate

2.8

3.2

550

9.0

0.010

0.10

1.4

5.6

Peak ConcentratlonUtg/Ll

Total

5.6

5.5

670

14

0.11

1.6

2.1

14

Dissolved

2.6

1.1

36

0.tt0.10

1.5

0.13

5.8

Particulate

3.0

4.4

630

13

0.013

0.14

2.O

7.7

Metal

Arsenlc

Copper

lron

Lead

Mercury

MolyMenum

Silver

Tlnc


Westmin's operations at Myra Falls includes the H-W, Lynx, Myra and Price mines,

which contain recoverable quantities of copper, lead, zinc, gold, silver, and cadmium.

Open pit and underground mining of the Lynx deposit began in 1966 with milling

commencing at the 750 tpd Lynx mill. Since that time, mill capacity has been increased

several times to the current 4400 tpd. The mill process uses conventional crushing and

grinding followed by selective differential flotation where separate copper, lead and

zinc concentrates are Produced.

After concentrator startup in Decemb er 1966, tailings were discharged to a small

nearby pond; however, Westmin obtained a permit in 1967 to discharge tailings to the

bottom of Buttle l-ake. Clclones removed the sand-sized material in the underflow for

use as backfill in the underground mine. Under gravity, ryclone slime overflow, at

7 to \yVo solids, passed through seven vertical drop boxes to a tailing raft several

hundred feet off shore, and were discharged via a submerged outfall that extended

below the thermocline. Flocculant was used to assist solids to settle to the lake bottom,

which was about 35-50 metres deep at the discharge point.

The tailings originated from the zinc circuit after milling of the high grade copper-lead-

zinc ore and consisted of sand-sizecl and silt-sizecl silicate gangue minerals and residual

copper, iron, lead and zinc sulphides. Heavy metal concentrations in the tailings solids

ranged widely but averaged 7000, 1300 and 900 mg/kg for Zn, Cu and Pb respectively.

Lime (CaO) was the only reagent used in significant quantity; approximately l.0kg/t

ore was added to the tailings to raise the pH in the milling circuits and to enhance

coagulation in the thickening tanks (Eccles, 1977). During the initial six years of

operation, with ore from the Lynx Mine, dissolution of heavy metals in the milling

circuit was minimal due to the high pH and extremely low solubility of metal sulphides.

In June lg7},the mill began limited production of lead ore (galena), and in early 1973,

this was increased when high-lead ore from Myra Falls came on-line. Because

production of a copper concentrate with a low lead content was required, cyanide was

used in the continuous copperleacl separation circuit. This resulted in substantially

increased levels of cyanide and clissolved copper in the effluent (at high pH, copper

complexes with cyanide). Consequently, Westmin introduced alkaline chlorination to

destroy residual cyanide and to precipitate dissolved copper in the tailings. Table 5-8

compares effluent before and after chlorination.

5-15


Jan.June 1973No Chlorlnatlon

Plant

Julv-Dec. 1973Chloilnation Plantln FullOperatlon

Maximum PermltLevels

TotalSolids

pH

Dissolved Copper

Dissolved Lead

Dissolved Zinc

Dissolved Sulphate

TotalCyanide

TotalChlorine

138,800 mg/L6.0 to 10.0

0.3 mg/L0.1 mg/Ls.0 mg/L

1000 mg/L0.5 mg/L

7.7

2.85 mg/L0.17 mg/L3.27 mg/L

4.72mq/L

70,500 mg/L9.74

0.09 mg/L0.08 mg/L0.43 mg/L380 mg/L0.18 mg/L

0.1 mg/L

t

Table 5-8

Comparison of Talllngs Enterlng Buttle LaksBefbre and After Commenclng Treatment

rNot recorded.Source: SENES/Knapp, 1981 (unpubllshed).

Some water quality and sediment chemistry clata for Buttle l¿ke are available'

Pedersen (1983) discussed the distribution of dissolvedZn, Cd, Cu, Mn and Fe in pore

waters extracted by centrifugation under nitrogen from four cores collected from both

tailings and natural sediments in the south basin of Buttle I-ake (Table 5-9)'

Concurrent with this investigation, dissolved Zn and Cu levels in the overlying lake

water were very high and it was speculated that remobilization of metals from the

deposited tailings was the cause. It was later realized that high metal concentrations in

Buttle l-ake were due to inflow from Myra Creek, which was draining a waste rock

dump containing sulphicle-rich wastes. Metal concentrations in the creek water

averaged 600 pg/LZn,40 pglLCu an<l 1.4 pglLCd, in July 1981 (8.C. Research, cited

in Pedersen 1983).

Although the sediment cores exhibited pore water metal concentrations which ranged

widely, it is clear from the data in Table 5-9 that diagenetic remobilízation of metals

was not substantial; in fact, metal concentrations in the tailings pore waters were much

lower than in the overlying lake water. In the interstitial waters of natural sediments

5-16


Table 5-9

Dissolved lnterstltial Metal Concentrations ln Buttle Lake Cores,Measured by Graphite Furnace AAS (Data from Pedersen 1983)

Sample Depth(cm)

Dlssolved Metal Concentratlon (Pg /LlMn FeCuZn cdCore

B1 0-33-66-1010-1414-1925-30354045-50

352027

31,3134

36,372623

<0.5<0.5<0.5<0.5<0.5<0.5<0.5<0.5

15.68.0,8.3

13.4,13.13.42.25.00.51.9

170130140170210

550,560't1490

241B

167.23.4231329

82 0-33€6-1010-1420-25354055€078-83

3434,34

f31923262930

<0.5<0.5<0.5<0.5<0.5<0.5<0.5<0.5

5.66.34.25.9

0.7,0.96.26.54.4

650800630

42623015

36131539

370,350650350

290,3002æ

B3 0-33€6-1010-1420-2530-3545-5060-65

33,316.4

78,7935285.8

42<3.0

2.7,2.1<0.5<0.5<0.5<0.5<0.5<0.5<0.5

1 3.2,13.12.O

2.04.7f .1

<0.50.5

<0.5

3190,31202740224023901610580520

7400,7330

4232233858

rt81

111

1550

B4 0-65-1010-1530-35

22,2021

7.4<3.0

<0.5<0.5<0.5<0.5

<0.5<0.5<0.5<0.5

1210026900

24900,2500044800

8700139009300

12900

l-ake bottom water54m depth, near 82

170 1 11

Lake bottom water87m depth, near 84

230 0.4 15

Sourcc: Pcdcr¡cn rnd Lothcr, 1988.

5-17


(core 84) the Zn concentration decreased rather sharply with depth, indicating that the

metal was being incorporated into an authigenic precipitate.

Pedersen (1983) concluded that several factors controlled these distributions. First,

oxidation of the tailings on the lake bottom was not ocsurring, presumably because the

rate of discharge was suffîcient to bury the deposits continuously and quickly, thus

minimizing the time of exposure to dissolved orygen in the bottom water. Because Zn

and Cu oryhydroxides are much more soluble than their sulphide counterparts,

oxidation of the detrital sulphides in the tailings would release metals to interstitial

solution. An example of this phenomenon is shown in Table 5-10. Water which was

allowed to accumulate by dewatering on the tops of two of the Buttle L¿ke cores

showed significantly higher Zn and Cu concentrations after the tailings had been

exposed to air for 8-11 h. The absence of any similar increase in dissolvedZn, Cu or Cd

concentrations with depth in the tailings cores (Table 5-9) confirmed that the tailings

were relatively unreactive on the lake bottom. There was no evidence that progressive

oxidation with concomitant metal release was occurring.

Table 5-10

Zn, Co anci Cu Gonceniraiions in Supern?taT^l_'YY?le-r lmmediatel'y-ÀfteiCoré Collection (A) and After ð-tt n of Partial Exposure of

the Tailings cinihe Top of the Cores to Alr (B)

Dissotved Metal Concentration (pg/Ll

CucdZnCore

(A)(B)

(A)(B)

B1

82

0.814.6,14.8

5.212.0,13.9

<0.5, <0.51.0,1.0

<0.5<0.5

22,20,2056,56

44'117

Source: Pedersen and Loshet, 1988.

Secondly, the absence of detectable dissolved Zn or Cu in the natural sediments

underþing the tailings in core B3 and the decreasingZn gradient in core 84 indicated

that these metals were diffusing into the natural sediments from the overlying

metal-rich lake water and were being removed from solution, presumably by

incorporation into authigenìc sulphide minerals. Support for this contention is given by

5-18


the associated increase of dissolved Mn and to a lesser extent, Fe (Table 5-9). Because

manganese and iron oxyhydroxides serve as preferred electron acceptors after orygen

and nitrate have been depleted (Chapter 3), their presence at high concentrations in

solution in the uppermost sample in core 84 is evidence that anoxic conditions are

established at shallow depths in the natural sediments. Iron sulphide precipitation in

this facies is of course limited by the scarcity of reducible sulphate, but is nevertheless

indicatecl by the decreased Fe concentration at depth in 84 relative to the steadily

increasing dissolved Mn level.

The Buttle Lake study demonstrated clearly the tailings were not diagenetically reactive

while deposition was proceeding at a high rate. In this type of example, the key factor

which mitigates against release of metals is the high sedimentation rate, which places a

strong limit on the amount of oxygen which can diffuse into the tailings from bottom

water to support oxidation of detrital sulphide minerals. As the data in Table 5-10

demonstrate, oxidation with associated metal release can occur quite rapidly in the

presence of high oxygen concentrations. Therefore, metal remobilization from the

lacustrine tailings could occur in the period following cessation of discharge but

preceding burial by subsequent natural sedimentation. In the south basin of Buttle

[-ake, the natural seclimentation rate measurecl at a representative site using 2l0Pb data

is ab<rut Zmmfyr (Table5-11). In the same core the solicl-phase Mn concentration

clecreases sharply between 0 ancl 3 cm clepth (Table 5-11), indicating solubilization at

depths below 1 cm. This distribution is consistent with the pore water data in core 84.

Assuming that the data from both cores are representative of the natural sediments in

the basin in general, then it can be suggested logically that the tailings would become

covered with a veneer of natural sediments, anoxic at a shallow depth, within 15 to

20 years following the cessatíon of discharge. At that point the metal-rich deposit

would, in effect, become chemically sealed with respect to upward diffusion of metals

from the previously oxidized, but now buried tailings.

Deniseger et al. (19SS) have recently reviewed the effects of increased heavy metal

levels in Buttle l¿ke water on inctigenous biota cluring the subaqueous discharge phase

of the mine operations, ancl have discussed the recovery of the biota during the period

19g3-19tt6. This three-year stucly periocl mostly postdates collection and treatment of

the leachate from the waste rock clump near Myra Creek (which commenced in 1983)

and cessation of lacustrine discharge in July, 1984. An abbreviated summary of their

report is presented below.

5-19

SELECTED CASE STUD¡ES

Table 5-11

2loPb (supported and unsupported) and Zn Goncentratlonsin Sediments from Core 88

Depth in Core (cm) etoro (dRm/O) zn (þslsl Mn (¡¡gls)

0-1

1-22-33-44-55-8

8-111 1-1515-2030-35

12.7 + 3.3

9.8 + 2.7

5.6 t 3.3

4.7 + 2.7

0.0 + 4.60.0 + 4.0

38004100

740184142128

f 0¿m0

4410

2170

2100

135 2030

Source: Pcderscn and Loshcr, 1988.

Metal levels in the lake reached approximately steady-state maximum concentrations in

1981. At that time, elevated levels of hepatic metallothionein measured in rainbow+-a"+ li.'ar. i- Þ"++l^ f -L^ ^-á i- l-Loo i-rra¡liofolrr rlnrrmcfraâm rr¡erp chmr¡n fnlt\rl¡L llvç¡ù lll uL¡LLlg HÃv cl¡u trt rcAvJ ll¡rr¡rvurqlv¡J svw¡^ù!¡vs^¡r

correlate well with mean dissolved zinc levels in the irrdividual basins (Roch et al.1982;

Roch and McCarter, 1984). Zinc concentrations as high as 440 ttgLr were measured in

the south basin of the lake during that period.

Major changes were observecl in the zoo- and phytoplankton populations as metal

loadings increased, including decreasing numbers of cladocerans and calanoid

copepods, and an altered phytoplankton species composition. By the carly 1980's,

zooplankton species such as Leptodora kindtü, Holopedium gíbberellum, Polyphemus

pediculus anrJ Daphmø spp.were essentially absent. By the late summer of 1985, when

metal levels had fallen by a factor of roughly 4-5 from their peak in 1981, these species

had reappeared" By 1986, species composition and diversity were similar to those

sampled in 1966-67 (Whately, 7969, cited in Deniseger et al., 1988). This improvement

is commensurate with falling metal levels in the lake, although it should be noted that

contrary to expectations, dissolved metal concentrations have yet to decline to their pre-

mine levels.

Similar changes were noted in the phytoplankton community. The dominant

association of Rhizosolenia/Asteríonettaftabellaría/Ceratíum/Perídiníum seen in

5-20


1966-67 was significantly different by 1980 (Deniseger et al', 19SS)' Rhizosolenia

eríensís,for example, essentially disappeared from the south basin waters where metal

concentrations were highest. This species remained scarce until late 1983. Massive

blooms of R. eriensrs persisted from mid-1984 to mid-1985, possibly in response to an

observed increase in the N:p ratio which may have accrued from increased discharge to

the lake of explosives residues, sewage anrJ eroclecl soil as the mine expande<J i¡ the

mid-1980's. In addition, R. erieruis appears to be more tolerant to metals than other

species, and the 1gg4-g5 bloom may have reflected a transitory competitive advantage

while metal concentrations were falling but were still relatively high (Deniseger et al.,

1e88).

Decreasing hepatic metallothionein levels in fish and increasing diversity of plankton

demonstrate that biological stress has been significantly reduced in Buttle l-ake.

However, Deniseger et al. (1988) note that Cu and Cd levels in fish livers remain

relatively high. Although injection of dissolved metals into the lake from Myra Creek

has been much reduced since 1983, concentrations in creek water, and thus the lake,

have not yet reached background levels. The phytoplankton community will apparently

require further improvements in water quality hefore the species composition can

return to its pre-mine state.

5.1.1.6 Kootenay Lake (Comínco's Bluebell Mine)Ríondel (Pb, Zn" Ag)

The Bluebell mine at Riondel on Kootenay l-ake's east shore was for many years a

major high-grade lead-zinc operation which also produced significant silver, cadmium,

copper and gotd. Tailings containing up to LTVo zinc, in addition to arsenic (in

arsenopyrite, FeAsS) were discharged directly into the lake. Although no acid

generation testwork on the mine wastes was available, it is expected that these wastes

were acid generating since the ore was a massive sulphide deposit. During the 1930's

the Bluebell mine had several different owners, but Cominco operated the mine during

its maximum production period from 1952 until it closed in 1972 (Kuit, 1989;

Envir<¡nment Canada/IWD: Daley et al., 19tì1; Douglas, 19tì4). In the 1960's, Cominco,

and recently others have sampled deposited tailings and conducted bathymetric surveys

to evaluate the feasibility of further metal recovery. Most of the tailings are likely

dispersed over a wide area of the lake bottom, however, because the slopes are steep.

5 -21


Small scale mining activity began in the Kootenay l¿ke area in 1890 (Environment

Canada/IWD: Daley et al. 1981). Almost all mines and concentrators around

Kootenay L¿ke disposed of their waste rock, mill tailings and mine drainage water

(sometimes softened with polyphosphates) directly into the lake. In the Ainsworth and

Riondel areas, tailings and waste rock containing iron and arsenic minerals from lead

and zinc ores were also dumped into the lake.

At the Bluebell mine, Cominco mined 4.8 million tons of ore during the 20 year

operating period. Waste rock was dumped into the lake to build a breakwater used to

shelter a marina. Barges carrying concentrate from Riondel to Proctor also routinely

dumped concentrate into the lake when they were cleaned and swept. In addition, four

rail cars carrying high grade Pb and Zn concentrates from a Cominco barge spilled their

cargo into the bay near Riondel in the mid 1960's.

Because of the history of metal mining activity adjacent to Kootenay [ake, several

studies in the late 1970's attempted to gather data to assess the impact of waste

dumping on sediment geochemistry and the distribution of metals within lake waters

(Daley et al., 1981; Chamberlain and Pharo, 1981). Some key findings from the 1979-80

Kootenay l-ake study are outlined below.

Overall metal concentrations in lake water were found to be low and reflectecl

background concentrations. Concentrations of certain metals were slightly higher near

the centre of the lake, possibly due to wastes from the Pilot Bay smelter.

Sediment core profiles showed strong enrichments of Pb, Zn, Ag, Cd and As in surface

layers, directly indicating the presence of mine wastes in the lake. Comparative data

shown in Table 5-12 illustrate the magnitude of the enrichments relative to other lakes.

Metal concentrations in the flesh of fish from the lake were also low. These

observations are also consistent with the results of leaching experiments performed on

the enriched surface sediments which suggested that, despite the high concentrations,

the metals were not readily bioavailable. However, Pharo (1989) admitted that the

design of the extraction test procedure did not provide firm conclusions regarding

potential accumulation of heavy metals in the food chain. Evans and l-asenby (1983)

and Vanduyn-Henderson and L¿senby (1986), in contrast, suggested that there may

indeed be some transfer of metals from sediments into the food chain (I-asenby, 1989).

It appears that further work is warranted in order to confirm or refute this suggestion.

5 -22

Table 5-12

(=Fåt5å',åîl.,i É1Ì,Xii'Tf¡ËHT'Í?l:'3Ttr'.,3å1fr331tiåî"'1',gBåiËil""oTilfr"oil3/f"*""

FBAFBAFBAFBAFBBA FA

Kootenay lakeCore F

Lake ErieWestern Basin2

Lake-.'ltsfre

LakeMonona

L¡keM¡chioan

Lake

ConstancaElement(ppm)

æ6zffi1880

4.46.15

83505022

309.1

600

48612

441.3

51.1

1.2

f.630æ

48æ

15/0.1

4.621637

4418

19

0.318

CuPb7Jl

AS

Fe(%)

MnNiCoCrcdAsr¡(%)

12

96

26812492

494.6

51't.12

1.5

3.52.5

1

33

ô

6

4

g

12

11

75

¿lO

18

7

1.0 2.3 2.3

nÆ

256

11

0,9o.2

0.0s

95 2.5

58

42

602.43.2

0.48

72

É5

50 1.5

72.5

2o.24

1

25

44

85

22o.2

33

4

3452

380

2214

15

75145

317

54

77

11

0.ø

44Q

1æ

50

r530.68

0.8

3019

't24

55

500.21

o.2

3.21.9

11

4

13

æ

8212

1253.5

0.

2.O

4.517

5.512

13

0.140.6

0.04

]Forstner (1924.iallan and Brunsklll (1977).-Thls study.

Source: from Daley et al. (f981).


5.1.1.7 Kootenay Lake (Dragoon Resources Ltd., formeþ David Minerals)Ainsworth (Pb, Zn, Ag Au)

David Minerals, an underground operation with a 135 tpd flotation mill, deposited

untreated tailings with a pH of 9.1 directly into Kootenay I-ake. Although littleinformation has been located, it is known that liquid effluent quality guidelines were

not being met in 1982.

5.1.1.8 Pinchi Lake (Comínco Ltd.)

Cominco's Pinchi I¿ke mercury mine, located 15 miles from Fort St. James, has

undergone two periods of operation; the first during World War II, and subsequently

from 1964 to 1972.

Mineral processing included a roasting step to form a calcine that was discharged into

the lake; tailings were impounded. Since the mine opened before there were

environmental regulations, no pre-operational baseline or background data exists.

Although no studies were done on the calcine, mercury has a low sublimation

temperature (583'C) and roasting of mercury ores is known to be quite efficient.

Consequently, it is unlikely that substantial quantities of mercury remained in the

calcine that went into the lake. Additional information on the handling and disposal of

mine waste was not located.

Due to depressed mercury prices, operations were suspended in 1975. Cominco still

holds the property and there is on-going water quality work on downstream receiving

waters to monitor existing mine and mill waters for mercury (Kuit, 1989; Canadian

Mines Handbook, 1988-89).

5.1.1.9 St. Mary's Ríver/Kootenay River (Sullivan Mine - Cominco Ltd.)

A survey of the St. Mary's River - Kootenay River system was conducted during 1965

and 1966 to evaluate the impact of wastes from Cominco's Kimberley Mine and

fertilizer plant on fish and other aquatic organisms. l-arge quantities of iron and acid

wastes were being discharged directly from the mine operations into upper Mark Creek,

and the Marysville fertilizer plant routinely released gJ¡psum into lower Mark Creek,

and ultimately the St. Mary's River. In addition, the Sullivan concentrator-

impoundment wastes flowed into Cow Creek, which emptied into the St. Mary's River

5 -24

SELECTED CASE STUD¡ES

(Cominco Amendments to Pollution Control Permit PE-189, November 23, 1982).

Deleterious effects of these effluents on the biota were suspected but had not been

extensively researched. The results of the Sinclair survey are summarized below

(Sinclair, L966).

Although the water quality of the St. Mary's River was excellent above Mark Creek,

extremely toxic con<titions existed in the St. Mary's River at a station below Mark

Creek. Fish exposed to these conditions died within 48 hours. Similarly, there was an

absence of benthic fauna in Mark Creek below the Sullivan operations, and in the

St. Mary's River below Mark Creek when the benthic fauna upstream of contamination

sources was abundant and healthy. Bottom fauna were extremely sparse as far as 15 km

downstream of the confluence of Mark Creek with the St. Mary's River and it appeared

that mine effluents had killed fish food organisms and greatly reduced native

populations of cutthroat and rainbow trout, mountain whitefish and Dolly Varden char.

Even 40 km downstream from the Cominco plant in the Kootenay River, the important

food organisms were significantly diminished although the large dilution flows in the

Kootenay had reduced toxicitY.

Water and sedíment sampling indicated high quantities of gypsum and other foreign

materials in the St. Mary's River, with gypsum deposits 10-14 cm deep in the river

bottom at Wycliffe. High concentrations of lead, zinc, and fluorides, all extremely toxic

to fish, were also noted, along with significant turbidity and discoloration.

5.1.1.10 Summit Lake (Scottie Gold Mines Ltd.)

Summit I¿ke is a 180 tonne per day underground gold mine. The mine and mill are

located at Summit Lake, approximately 32 km north of Stewart, B.C.

The lake is an ice-dammed glacial body of water lying at the head of Salmon Glacier at

an elevation of approximately 823 m. The entire volume of the lake has discharged

under Salmon Glacier into Salmon River numerous times. As the ice dam has

retreated and thinned, the lake discharge has become an annual fall event.

The tailings pond was constructed below the historical high-water table, hence it is

inundated with lake water during the latter stages of the fill cycle (May-August). At

other times of the year, the tailings remain in a saturated state since the groundwater is

essentially at the surface.

5 -25


The tailings were demonstrated to be acid generating, which reflected the high pyrite

ancl pyrrhotite composition of the ore. The sulphide content of the tailings was not

determined, however, tests on the waste rock showed the rock to be acid consuming'

All waste rock was utilized for fill and road construction.

Typical deposited tailings characteristics are shown in Table 5-13 for three locations'

No recent water quality data were found; however 1979 water quality are presented in

Table 5-14.

Table 5-13

DepositedTailingsCharacteristicslnSummitLake

Note: All values erPrclscd ln ug/g.

The mine and surface streams were of relatively good quality; Summit L-ake was typical

of a glacial lake with high turbídity and solids.

30 m below break 6O m below breakNear splgot

Location

AlumlnumArsenlcBariumBerylliumCalciumCadmiumCobaltChromiumCopperlronMercuryMagnesiumManganeseMdyMenumSodiumNickelPhosphorusLeadSiliconTinStrontiumTitaniumVanadiumTinc

23600162019.5<o.2

3530022.i20517.9691

163000<0.008

181001590

8.660<3

1080750890<244

1670167

1750

19100300018.7<0.2

31300ct.o35413.3

1 100

244000<0.008r4700

14Ít0<0.8

707

10901340780<2

41.71350

1532960

24æ063730.2<0.2

ar200ô^ô135

22.3908

152000<0.008

18200172034.8120<3

1 190759

1850<284

1530179

1710

AIAsBaBeCacdCoCrCuFeHgMgMnMoNaN¡

P

PbSiSnSrTiVZn

5-26


Creek Water atSummit Lake CamP Summit LakeMineWater

pH

Cond uctlvltY (mlcromhos/cm)

Turbidity (JTU)

Total Suspended Solids (mg/L)

Dissolved Sulphates SOo (mO/L)

Dissolved lron Fe (mg/L)

Dissolved Cadmium Cd (mg/L)

Dissolved Copper Cu (mg/L)

Dissolved Lead Pb (mg/L)

Dissolved Zinc Zn (mg/L)

Dissolved AntimonY Sb (mg/L)

Dissolved Arsenic As (mg/L)

Dissolved Cobalt Co (mg/L)

TotalMercury Hg (mg/L)

7.90

223.

2.8

12.2

44.0

<0.030

<0.001

<0.00f

<0.001

0.002

0.024

0.009

<0.005

<0.0002

6.45

16.4

2.1

0.6

6.0

<0.030

<0.001

<0.001

<0.001

0.004

<0.003

<0.001

<0.005

<0.0002

7.10

56.6

380.

35.2

10.0

0.047

<0.001

<0.001

<0.001

0.002

<0.003

<0.001

<0.005

<0.0002

Table 5'14

water Quality at scottie Gold, 1979

5.I.2 Other Canadian Mine Sites

5.1.2.1 Fox Lake (Fartq and Shenídon Mines, Shenitt Gordon Mines Ltd')

Sherritt Gordon Mines Ltd. intermittently operated the Fox I¿ke copper-zinc-

gold-silver mine from 1930 to Lg52. On average, 3,000 tpd were milled during this

period. Although it is known that both on-land and subaqueous tailings disposal into

Fox l-ake were used, subaqueous disposal is thought to have been practiced only during

the last few years of operation (1949 to 1952). Information on the quantity and mineral

composition of the tailings is also scarce. Kennedy and Hawthorne (1987) studied the

high sulphide tailings which were disposed on-land and those which were submerged,

and attempted to compare the identity and character of the authigenic minerals, the

cxtent <lf sulphide mineral oxiclati<ln and the identity ancl composition of alterecl and

unaltered silicate minerals.

5 -27


Little is known about the tailings sampling procedures and techniques which were used

for this study; apparently, field notes were not taken. From the Farley and Sherridon

tailings disposed on-land, five samples, including hardpan ancl material which was

poorly consolidated and thoroughly oxidized, were selected and examined. Of the

submerged tailings, 22 samples from a single borehole (BH5) were recovered from the

tailings submerged in Fox I-ake. Both tailings samples were screened into their

>200and -200mesh fractions. A magnetic component was also recovered from

selected sand fractions. Unfortunately, no pore water samples were recovered from the

submerged tailings.

Cold mount epo{y impregnation was used to prepare the samples as polished thin

sections. Samples were subsequently analyzed using optical and electron probe analysis

of minerals within the thin sections and optical examination under transmitted and

reflected light microscopy. Powder X-ray diffraction (XRD) was also used to examine

the different size fractions of the oxidizecl land tailings samples, as well as those of the

submerged tailings samples. Where XRD was not successful due to sample size, a

Gandolfi camera was used to obtain powder diffraction patterns.

Tables 5-15 and 5-16 illustrate the general mineralory and texture of both the oxidized

land-disposed tailings and the submerged deposits in Fox [-ake. Silicate minerals

comprised 65-75Vo of both the land ancl underwater tailings. Phlogopite and biotite

mica were predominant in the underwater tailings. The land tailings contained quartz

(25-30Vo), albite (plagioclase) and potassium feldspars, and small amounts of

magnesium-rich chlorite, hornblende, biotite, phlogopite and muscovite.

In the land-disposed tailings, gJ¡psum (CaSOo.2HzO), jarosite [KFer(SOJ2OHó],

goethite (FeOOH) ancl analcite, Na(AISirO6)-HrO are the principal authigenic

minerals present in the hardpan. Authigenic minerals are those derived from the

oxiclation of sulphides and percolation of acicl-bearing solutions" Goethite and jarosite

are the main cementing ingredients. Hardpans in land tailings are an indication of

incomplete oxidation of relict sulphides (pyrite and pyrrhotite); thoroughly oxidized

tailings contain no sulphides, due to their clissolution into percolating porewaters, and

as a result are poorly cemented.

On-land tailings rarely contained the mica-type minerals biotite and phlogopite. This is

attributed to the oxidation of sulphicle minerals and associated generation of sulphuric

acid, which reacts with the mica-type minerals, altering them to finely interlayered clay

5 -28

Table 5-15

Description of On-Land Tailings from Sherridon'

Note thc lndþadon ol th. rclâüyc proporüon of rulphlder end euthlgenlc mlneral¡ rnd that all authlgenlc mlnerals are hydrated. Sample No. BH7-S2 rlso contaln¡ enalclte'

l,la(AlSlrOJ' H2O, wh¡ch l! Probably .uthlgenlc.

lFrom Kennedy and Hawttrorno 1987.

On-L¡nd Tailings, Oxidized Samples

Sulphide Minerals SuGolourDescription

Authigenicand Oxide Minerals

TP12Æ Hardpan: wellcemented Deep reddish brown <8%Pyrite > Pyrrhotite

Goethite > Jarosite > GYPsum

TP7-28 Hardpan: wellcemented taminated: light brown andyellowish brown

<5o/o

Pyrite > PyrrhotÌteGoethite >(trace Lepirlocrocite)

BH7€2 Hardpan: mderately towell cemented, alongmargin of more oxidizedmater¡al

Reddish brown (also with moreoxidized, yellowish brownmaterial)

<15o/o

Pyrtte > PyrrhotiteJarosite >(trace) (SulphirJes > Authigenicminerals)

TP-BH2 Poorly cemented, friable Bright, yellowish brown toorange brown

None Lepidocrocite > Jarosite > GYPsum

Tablre 5-16

Description of Bore Hole Five (BHS), Sherridon Tailings, Beneath Fox Lake'

Authlgenlc mlner¡l¡ are llsted ln approxlmately order ol abundance'Anhydrltc 13 repoÌted from Cu-Zn mlnes ln the Lynn Lake area, and may not be an authlgenlc phase.rFrom Kennedy and Hawthorne 1987'

Length Depth GrainSize Comments

Extent ofSulphideOxidation

AuthigenicMineralsAmple (inches) (feet

BH5-15-25-35-4

0€6-12

't2-1616-20

(0) SandSandSandSand

Piner needles, plant remains:LooseLooseLoose

3540%20-25%15-20øA

10%

Jarosite,Goethite,Gypsum

BH5-55€5-75€

20-2626-3232-383842,5

SandSandSandSand

Compacted, denseConrpacted, denseConrpacted, denseConrpacted, dense

30-35%15-20%

25%25-35%

Goethite,Jarosite,Gypsum

BH5-95-105-115-12

42.54€..5¿18.5-il.5

54.5€0.560.5€6.5

SandSandFine sandFine sand

Compacted, denseConrpacted, denseCompacted, denseConrpacted, dense

25-30%30%25%25%

Goethite,Jarosite,Gypsum

BHs-l35-14

66.5-72.572.5-78.5

Sand,Sand,

siltsilt

l-aminated lnterlayered sand andsitt; appreciable sphalerite,trace chalcopyriteCompacted, denseConrpacted, dense

sand 15-20%silt <10%

5-155-16

78.5€4.584.5-90.5

Sard, fine sandFine sand

15-20%GypsumJarosite,Goethite

BHs-175-185-195-20

90.5-96.596.5-1 1 1.5

1 11.5-116.51 16.5-121 .5

Silt, fine sandSiltS¡It

silt

Conrpacted, denseConrpacted, denseConrpacted, denseConnpacted, dense

<5-10% Anhydrite*(trace Jarosite,Goethite)

BH5-215-22

121.5-126.5 (21)21-26

Silt, mudSilt, mud

Lake bottom sediment, no sulphideLakra bottom sediment, no sulphide

N.A.N.A. Gypsum


minerals, including chlorite. Kennedy and Hawthorn (1987) suggest that the micas act

as sulphuric acid buffers, liberating Fe3+ and K+ ions and promoting the formation of

goethite and jarosite.

In contrast, submerged tailings exhibit abundant biotite and phlogopite, indicating far

less acid-induced alteration due to sulphide oxidation. Oxidation of sulphides is

advanced in the sandy-textured tailings near the sediment surface, and has consumed an

estimated 20-35Vo of the pyrite and pyrrhotite. Although this is a lesser degree of

oxidation than that of the land tailings, such reactivity was unanticipated in view of the

generally high water quality in Fox l¿ke. The degree of sulphide oxidation correlated

directly with grain size; only minor oxidation was noted in the silty, less porous fraction.

Sulphide grains often displayed prominent oxidized coatings of cryptocrystalline

goethite.

Goethite, jarosite and gypsum are ubiquitous authigenic minerals in the submerged

deposits. Kennedy and Hawthorn (1987) note that phlogopite, Mg-biotite and

muscovite have reacted extensively with the acid generated during sulphide oxidation.

Pertinent reactions indicating the formation of the principal alteration products are

shown in Table 5-17. Heavily altered phlogopite and biotite are replaced by very fine-

grained aggregates of chlorite, montmorillonite, quartz, illite and kaolinite, with

intersitial jarosite. Virtually all the mica grains have been altered to some degree,

which is manifest by potassium depletions.

Kennedy and Hawthorn (1987) suggest that the observed mineral distributions, which

clearly indicate extensive oxidation, reflect the percolation of orygenated lake water

through the sandy tailings after deposition. However, it is not known if oxidation is still

proceeding. It is also not clear whether or not the tailings were initially stored on land

and subsequently submerged or originally deposited directly in the lake. Kennedy and

Hawthorn make the interestíng observation that the large volume increase associated

with the formation of hydrated ferric oxides and jarosite from the oxidation of sulphides

must eventually restrict permeability. They speculate that these reactions have the

potential to seal the deposits with time and eventually inhibit continued oxidation.

5.1.2.2 Ganow Lake (Polaris Mine, Cominco)

The PolarisPb/Zn mine on Little Cornwallis Island, N.W.T., commenced discharging

tailíngs into proximal Garrow l-ake when the mine opened in 1981. The lake is

5-31


Tabte 5-17

Mineral Reactions Assoclated wlth the Oxldatlon of Sulphlde Minerals lnReactive Mlne Talllngs

tK ¡" t"k"n up by other mtnerat reacllons, or goos lnto ¡olutlon.

Source: Kennedy and Hawthorne (1987).

approximately 3 km long by 2 km wide with a depth of 46 m, and its surface is only 7 m

above sea level. Runoff enters the lake from a relatively small watershed. Surrounding

elevated terrain shelters it somewhat from the prevailing wincls. Limnological studies in

the 1970's establishecl that Garrow L¿ke is meromictic, with a slightly brackish aerobic

zone overlying anoxic HrS-bearing brine.

The lake is stratified into three distinct hydrological regimes (Bohn et al., 1981; Kuit

and Gowans, 1982). The brackish surface layer (epilimnion) extends to a depth of 13 m

and is subject to inflows and surface mixing. The transition zone (metalimnion) from

l3to20m depth exhibits decreasing dissolved orygen, increasing salinity, and is

influenced only to a limited degree by wind-induced mixing. Permanently stagnant'

highly-saline bottom water, reaching about 9 7ooS, extends beneath the transition zone

(the hypolimnion) to a depth of 46 m. Characteristic temperature and salinity profiles

OXIDATION OF SULPHIDE MINERALS

(1) Pyrite: FeS, + 3'50. * Hp

(2) Pyrrhotite: FerS. + 15'50. + H.O

(3) 4FeSOo + 2H.SO. + O

(4) Fe.(SO.). + 4H.O

FeSO. + H,SO.

7FoSO. + H.SO.

eFe.(SO.). + 2H"O

2FgO(OH) + 3H-SO.(goctÈite ôr lepidocröche)

REACTION OF SULPHURIC ACID WITH SILICATE

(s) Mg.Brotrte 'äir.";:i:ä.;Jjffi, + 12rH.or + ro.r --->

2[MgrFe.Al..u(sio.uAl..ooJ(oH),.] + f StsloJ + 2[KFe.(SO1)2OH.] + 3KrO*

(6) Muscovite --> Jaroslte + Kaolinite

õtv ^t /ê¡ ^l f'ì \/rìU\ I r lofFaQô I ¿ 9ôfl.l nl = 6fO I ---->¿lr\2¡ll4\9r6^rrvrl\v¡ rr4¡ , rÉu vvv4t , -r'2.

3[Ald(S¡1o,0) (oH).] + 4[KFer(So.)r(oH)r]

2

5 -32


are presented in Figure 5-5, and distributions of selected chemical parameters are

shown in Figure 5-6.

The ore at polaris consists of sphalerite and galena in a matrix of marcasite, calcite and

dolomite. The tailings are therefore enriched in Pb and Zn and contain high

concentrations of FeSr. Due to the nature of the ore (i.e. massive sulphide deposit) it is

expected that the tailings are acid generating; however, the exact sulphide content was

not determined.

It was originally anticipated that the relatively high dissolved sulphide concentrations in

the bottom water would render the lake an ideal repository for sulphide-bearing

tailings, and that the release of Pb andZnto solution would be negligible, providing the

tailings were deposited in the anoxic hypolimnion. However, complications arose

during testing of simulated discharges. First, in order to minimize pluming and areal

dispersion of the tailings, and avoid contamination of the oxic epilimnion and

consequent harm to the indigenous biota, the behaviour of discharges having different

densities was examined. Tests which simulated the direct discharge of concentrator

tailings at 32Vo solids pulp density predicted that substantial pluming and dispersion

would occur. In addition, and unexpectedly, this particular mixture consumed copious

quantities of HrS from the receiving solution, which permitted the dissolution of Pb and

Zn sulphides (Kuit and Gowans, 1982). Apparently, the consumption reflected the

adsorption of HS- onto the surfaces of iron-bearing minerals. The extent of sulphide

depletion was shown to be inversely proportional to relative particle size and

proportional to the total iron content. As Kuit and Gowans (1982) note, these

phenomena demonstrate that minimizing dispersion is a critical factor in limiting the

the post-depositional reactivity of tailings solids.

Following further tests, thickener underflow at 60Vo solids pulp density was chosen as

the optimum mixture for discharge to the lake. Pluming of this slurry was minimal and

sulphide adsorption was negligible, being restricted to a thin Iayer on the surface of the

settled deposits.

In actual practice, the tailings discharge points are locate<l well below the halocline in

Garrow L¿ke. Pulp density of the tailings during 1,982-83 was maintained at an average

55% solicls. Weekly monitoring of the water column for the initial L8 months of

operation showed little or no change in water quality in the epilimnion; metal

concentrations remained well below maximum values stipulated by the discharge

5-33

IA

IA

ro

\o

\ a-ao

A Solinity

o:-.A

IA

IA

Tempero lure

IA

IA

-4 -2 0 2 4

Temperolure ("C

68lot2t4 16 18202224

to

rR

Eo.(¡¿

o

:^

?E

40

AC

20 30 50 60

( ports per thousond )

40

Solinily

70 80 90 rooo ro

Figure 5-5 Garro¡ Lake - Salinity and Temperature Prof¡les (Kuit and Gowans, 1982)

(@

o 30 60

o

40

50

o

o

to 20

75 o 50 o o.5 o 15 -25

Figure 5-6 Typlcal Physlcochemlcal stratlflcatlon of Garrow lake by water Depth

(Ouellet and Page. 1988)

20Depth

(m)

-t2 oo 5

m

c

M

Ct lg/L)I

I

\

\\

@/oo)

r860Mn, Fe(m9lL)

Tolo I

Phosphorus

(mg /L)

Solinily (O/O

4

I lurbidiry(JTU)

H S (moll)2-

O" (mg/Ll

I ("c)

(@


license, as shown in Table 5-18. Continuing monitoring indicates that surface water

quality in the lake has not been compromised by its use as a receptacle for tailings

(W.J.Kuit, 1989, pers. comm.).

Table 5-18

Garrow lake Surface Water Quallty, 1982

Parameter License Maxlmum, mg/L Actual, mg/L

TotalCuTotalPbTotalZnTotalCN

.o7

.020.151.0

<.008.01

<.o2<0.1

Source: Kult ¡nd Gow¡n¡ (1982).

5.1.2.3 Mandy Lake at Mandy MineHudson Bay Mining and Smeltíng Ltd., Flín Flory Manítoba

Mandy Mine was located near Flin Flon, Manitoba. The ore vein averaged over 20Vo

copper as solid chalcopyrite and contained significant precious metals in lower grade

sulphides. From April 1943 to December 1944, the mine deposited tailings into Mandy

Lake, a shallow water body next to the property ( Hamilton and Fraser,1978).

Hudson Bay Mining and Smelting Ltd. (HBMS) conducted ¡vo sampling programs in

1975 and 1976. Investigation of predominantly pyrite tailings which were submerged in

shallow water (0.3 - 1 m deep) for 3?years was carried-out to determine the extent of

oxidation. The original chemical and mineralogical composition of the tailings is not

well known, but pyrite was the predominant iron sulphide mineral and the ore likely

had a sulphur content of ß-fi%o. I-evels of zinc and copper are thought to have been

appreciable"

Tailings samples (four underw,ater, two onshore) and water samples (four taken above

the submerged tailings samples, two taken in Mandy t-ake inlet and outlet streams)

were analyzed. Results were compared to analyses of land-disposed Cuprus Mine

tailings, located 11 km southeast of Flin Flon, because these also were high in sulphides

and had been exposed for the same length of time. Tables 5-19 and 5-20 summarize

and compare the analytical data (Hamilton and Fraser, 1978). Reactivity generally

s-36


correlates with the intensity of exposure to oxygen. Tailings with some degree of water

cover were found to have oxidized to a lesser degree as indicated by the iron oxide

content (Table 5-19) and the leaching behaviour (Table 5-20).

Table 5-19

Gomposltlon of Tailings from Mandy Mine and Cuprus Mine

Source: Ham¡lton ¡nd Fra¡cr (f 978).

Table 5-20

Soluble Constituents ln a 1:5 Distilled WaterLeach of Taltings from Mandy and Guprus M¡ne

Constituent (%)

Mandy MíneUndenrvater Shore

Cuprus MineTop Profile

Totallronlron as FerOaTota¡sulphurSulphlde sulphurSulphate sulphurAlumina (Al2Os)

Silica (SiOr)

CopperZincLead

CadmiumCalciumMagnesium

17.4

0.8

15.5

15.4

0.03

8.4

35.7

0.9'l

4.70

0.13

0.01

1.04

1.75

19.8

2.9

15.8

15.5

0.30

6.737.7

2.70

1.60

0.12<0.01

0.51

1.03

24.5

15.8

6.6

2.7

3.9

2.9

27.7

0.19

0.18

0.08<0.01

2.87

0.54

Constituent(mg/kg tailings)

Mandy MineUnderuater Shore

Cuprus MineTop Profile

lronCopperZincLeadCadmiumCalclumMagnesiumSulphatepHConductivity (mmho/cm)

0.90.25

15.50.350.05

14518

2N6.90.4

84.36.23

44f .050.28

2,590121

7,3404.92.45

8,530797218

1.270.57

2,360577

35,4002.57.20

Sourcc: Hamllton rnd Fra¡cr (19781.

5-37


Comparison of the quality of inlet and outlet waters to ancl from the lake (Table 5-21)

suggests that the submerged tailings had no significant influence on the dissolved metal

inventories during the 1975-76 study. Furthermore, animal species were present

universally throughout the non-affected and tailings-affected areas; no distinction was

evident between the two.

Table 5-21

Quality of lnlet and Outlet Water at Mandy Lake

Source: Hamllton and Fraser (1978).

Vegetation was well established on the water-covered deposit in the mid'1970's but was

absent on the subaerial tailings. Hamilton and Fraser (1973) noted that a layer of

decaying organic material up to 2.5 cm thick coverecl 90-95 o/o of the surface area of the

submerged dep6sit. Presumably, such layers enhance oxygen consumption and further

reduce the possibility of oxiclation of sulphides. This is consistent with studies by

Jackson (1978) on Schist [,ake, also near Flin Flon, where deliberate stimulation of

algal blooms through nutrient enrichment in a small lake serving as a disposal site for

toxic metal wastes causecl metals to be trapped and immobilized in the bottom

sediments as sulphides or metal-organic complexes. Fertilization with sewage or other

sources of available nutrients results in algal blooms which have been found to

concentrate metals, transporting these metals to the lake bottom when the algae die-off

and sink. Reducing conditions caused by putrefaction of algal organic matter tends to

further immobilize metals in the organic bottom muds. Moore and Sutherland (1981)

Sample Source

Above TailingsOutletlnletParameter

pHConductivity (mmho/cm)Copper (mg/L)Zinc (mg/L)Lead (mg/L)Cadmium (mg/L)lron (mg/L)Calcium (mg/L)f\r!annocirm lmo/Lì

\_ ' '9' *'

Sulphate (mg/L)

7.50.200.01

0.140.010.010.11

22.05.3

28.8

7.70.390.020.130.010.010.23

22.55.5

21.7

7.70.260.0f0.200.010.010.05

21.O

5.6

5-38


similarly found that heary metal releases into Great Bear l¿ke, Northwest Territories,

had become bound in the organic sediments, resulting in little impact on metal

concentrations in the overlying water and in fish flesh.

In summary, few, if any, detrimental effects on water quality were observed in Mandy

I-ake some three decades after deposition of the tailings, and the submerged sulphide-

rich material appeared to be essentially unreactive. In contrast, the land-based tailings

were heavily altered ancl had apparently undergone severe oxidation, presumably with

concomitant acid production. There are, unfortunately, no data available which would

permit evaluation of the impact on the lake of the submerged tailings in the immediate

and short terms following emplacement. However, HBMS is planning to submit a

proposal that Mandy Lake be considered as a target for field tests under CANMETs

Mine Environmental Neutral Drainage (MEND) program (Musial, 1989).

5.1.2.4 Anderson Lake (Hudson Bay Mining and Smelting Ltd')

Hudson Bay Mining and Smelting's Snow Lake mill, rated at 3450 m/tpd, receives

various copper-lead-zinc ores from seven Snow l-ake area underground mines (Fraser,

1989; Canadian Mines Handbook, 1988-89), and discharges tailings to proximal, and

shallow, Anderson l-ake. During the last decade, about 0.75 million tons of tailings

have been deposited in the lake annually.

Disposal locations on the lake are varied seasonally, with the more shallow depths

(3 m) being selected in summer and greater depths (average 6 m) used in winter.

Discharge is through a 30 cm floating Sclair pipe which can be moved to distribute the

tailings from two discharge points in order to keep a water cover of 0.6 m. I-ake bottom

contours are measured yearly as part of planning the seasonal discharge.

There are no {etailed mineral<lgic or compositional analyses of the tailings, which vary

with the composite ore feed from the clifferent mines. þrite and pyrrhotite are the

primary sulphide gangue minerals in the Anderson I-ake chlorite schist host rock.

Sulphur content is 20-25o/o.

A pre-operational environmental assessment was carried out and impacted receiving

waters undergo further assessment every three years. Periodic water quality monitoring

is done for metals, pH and suspended solids; federal effluent guidelines are reportedly

being met (Fraser, 1989).

5-39


Anderson Lake is marshy and eutrophic, floored with a soft organic ooze, and

historically has contained very few fish. Tailings initially deposited into the lake

maintained a 25-30 " angle of repose, and did not readily disperse due to the very

quiescent lake conditions and the short drop from the discharge outfall. Initially, some

mine waste and tailings built up above the lake surface and had to be moved in order to

keep the deposits submerged

5.1.3 InternationalMines

5.1.3.1 Silver Bøy, Mínnesota, on Lake Superior (Resere Miníng Co.)

A brief summary of the Reserve Mining Company case study is included here to

demonstrate that chemical behaviour is not the only aspect of lacustrine tailings

discharge which has important implications to environmental quality.

Following a 1950's technical breakthrough that led to the economic processing of low

grade, siliceous and abrasive magnetic taconite iron ores, the Reserve Mining Co. of

Silver Bay, Minnesota startecl beneficiating iron ore at a plant northeast of Duluth on

Lake Superior" The company was permittecl to discharge a large quantity of tailings

directly into l-ake Superior and built a concentrator that would process 88,500Itpd of

crude feed (gra<ting24-25Vo iron) and produce 59,000 ltpd of tailings. The tailings were

composed of about 50Vo quartz,43Vo other silicate minerals (including cummingtonite,

an asbestiform mineral), and 7Vo magnetite (Oxberry et al., 1978).

Although its effectiveness had not been proven, lake tailings disposal was initially

accomplished by allowing the tailings to flow as a density current downslope to an

offshore trough which ranged in depth from 200 to 300 m. Studies were initiated at the

same time to predict the properties of the tailings suspension in lake water, and to

predict settling behaviour (Engineering and Mining Journal 7972, 1976). When the

discharge entered the lake, coarser particles quickly settled to form a delta (Temple,

1980), and the fines flowecl in density currents tlown the delta face to form prodelta

deposits. However, a portion of the fine-particle suspension was later found to escape

the density current and under certain conditions fine particles became widely dispersed.

Eventually, tailings settled over more than 2500 km2 of the lake floor. The tailings were

found to affect organisms in direct toxicity tests and to have inhibited algal

5-40


photosynthesis (Shaumburg, L976). Atthough the tailings released soluble components

when exposed to lake water, no heavy metal releases were reported.

In 1969, the tailings were determined to be polluting l¿ke Superior waters, and the

Reserve Mining Co. was charged by the U.S. government. What began as a water

pollution abatement case became, after the discovery of asbestos-type fibers in I-ake

Superior, a public health impact issue concerning asbestiform particles discharged into

the air and water (Durham and Pang, 1976). Eventually, in 1978, Reserve was forced to

use onland tailings disposal, at an estimated cost of U.S. $250 million.

5.1.4 Flooded Pits and Shafts

5.1.4.1 Endako Mines Division (Placer'Dome Inc.)

There are two open pits at Endako's molybdenum operation near Fraser l-ake in north-

central British Columbia; the initial Endako pit, which was closed in June 1982 due to

depressed molybdenum prices, and the Denak pit. The Endako pit was allowed to

flood in 1982.

Some recent water quality monitoring data from the Endako pit were provided by

B.C. MOEP (Roberts, 1989) and are given inTable 5-22:

Table 5-22

Water Quality in Endako Pit

ConstituentAnalysis

(ms/Q

sooCucN"Mo

-2 Ranges 896-4250.001, 0.002, 0.003, 0.01

n.d.15

Discharge from the Denak pit to an intermittent creek that flows to the Endako River

or via a settling pond to Watkins Creek and thence to Francois I¿ke contains about

1.5 mg/L Mo, about 500-fold higher than the B.C. Water Quality Criterion for wildlife

(0.05 mg/L). Concentrations as high as 86 mg/L (J.8. Brodie, unpublished report,

5-41


1986) have been measured. Such high concentrations in the effluent from both the

Denak and Endako pits are of current concern to the Waste Management Branch, and

have serious implications for the eventual abandonment of the mine, particularly in

view of the probability that the pits will be purposely or passively flooded once the mine

closes.

Studies are underway which will address the impact to date of effluent discharges on

receiving waters, soils, vegetation, and wildlife in the mine vicinity. At the time of

writing, no results from these investigations are available. Mitigation of the release of

Mo to pit waters will need to be considered as part of any abandonment plan, as

discussed in Chapter 6.

5.1.4.2 Equity Silver Mines Ltd. (Placer-Dome Inc.)

Equity Silver Mines Ltd. (ESML) is locatecl near Houston, B.C. in the Bulkley River

drainage system. Acidic drainage containing dissolved Cu, Zn and Fe at the mine's

50 million tonne waste rock dump results from oxidation of pyrite, and was first noted

just one year following the mine's opening in 1981. The acid drainage flowed into the

Bulkley River and threatened to degrade valuable fish habitat and drinking water.

ESML subsequently initiated a reclamation and revegetation program, and constructed

a series of collection ditches which direct acid mine drainage (AMD) to an on-site

treatment plant. Approximately 800,000 m3/yr of acid drainage is mixed with lime to

raise the pH to 8.5, which fosters the precipitation of metal hydroxides, and the solution

is then discharged to a settling pond where a high proportion of the previously-dissolved

metal inventory settles out as a calcium sulphate/metal hydroxide sludge. The resulting

clear supernatant is discharged to the environment; according to ESML s

Decommisioning and Closure Plan (1988), such discharges are carefully monitored and

regulated to ensure minimal impact, and they comply with established water quality

guidelines. Representative water quality data are shown in Table 5-23 (Gallinger,

1988). The metal-rich sludge from the settling ponds is currently mixed with tailings in

a ratio of 1:10 and added to the regular tailings ponds.

Several alternatives currently exist for long-term pit abandonment, AMD treatment and

associated sludge disposal. The Main Zone and Waterline open pits will be allowed to

fill and form a lake. The tailings pond will become a shallow lake, submerged to a

5-42


Values ln mg/LpH Acidity SOo Cu(d) zn(dl Fe(d)

Raw A.M.D.

Treated

Permitted

2.35

7.80

6.5-8.5

10,000

NIL

8,500

1,600

120

0.01

0.05

80

0.04

o.2

800

0.03

0.3

Table 5-23

A.M.D. Treatment StatisticsESML's Water Quality Data (Gallinger, 1988)

ma,ximum depth of l-2 m; submergence is expected to lessen susceptibility of tailings to

oxidation.

The Southern Tail pit has already been flooded, following partial backfilling with

2.5 million cubic metres of waste rock; acid generation under the water in the pit is

expected to be obviated. Ongoing mitigation of the waste-dump AMD will produce

approximately 80,000 rÑ (LVo solids) of sediment sludge annually, which may be

deposited in an existing diversion pond basin (20 year minimum capacity) or in the

Main Pit (on the order of 200 year capacity). The latter option has the benefit of long

Iifetime, but it may be incompatible with the current plan to flood the Main ZonePit

f<lllowing abandonment. Concern for the potential generation of AMD from the

exposed pit walls cluring the six-year filling time is expressed in ESML's Closure Plan

(1938). Should a low pH result (and possibly persist), the potential for dissolution of

hydroxide precipitates in the sludge exists. Should this become the case, then the

anticipated establishment of aquatic life in the flooded pit will be compromised.

Provision will be made to route Main Zone overflow to the AMD treatment system

should it become necessary to treat acidic overflow water from the newly-created lake.

Such a circumstance will require alternate arrangements for the disposal of the sludge

to be made, possibly in proximal landfills. However, indications from other sites, such

as in Mandy lake (Section 5.1.2.3), imply that eventual submergence of sulphides in the

pit walls should inhibit oxidation and concomitant acid production. Such a result would

obviously be beneficial, and would limit the period of concern to the six-year filling

time.

5 - ¿lÍl


Other mine operators, for example, HBMS (Typliski and Labarre, 1980) have

successfully used lime neutralization to clarify mine waste effluents and recover metals,

for a net economic return. Presumably this is not an economically viable option for

ESMI4 as it is unremarked in the Decommisioning and Closure Plan (1988)'

5.1.4.3 Phoenk copper Mine fformerþ with Granby Mining co.)Decommíssioned OPen-Pit

Kalmet (1989) reports that the Phoenix open pit, abandoncd in the late 1970's, is

flooded and contains very clear water. Since the host rock is limestone, thc acid

generating potential is low. The most recent water samples werc collected from the pit

in 1984 (Jarman, 19S9) and were analyzed for metals an{ nutrients (Table 5'24)' At

that time, the water contained concentrations of nitrate which exceeded drinking water

guidelines, low concentrations of lead and relatively high levels of Cu, Mo and Fe' The

source of the metals is not clear.

Following cessation of mining, the tailings impoundment was breached by cutting a

channer into the dam to permit the crraining of ponded water. This drainage is diverted

via a perimeter ditch to Twin creek, the water quality of which is monitored (Kalmet,

1989; Jarman, 1939). Data collectecl <luring the course of monitoring downstream

receiving waters between 1973 and 1984 indicate that copper and sulphate

concentrations decreased significantly during that 11 year period (Table 5'25)'

Reclamation of the tailings impoundment is now complete.

5.1.4.4 Ciry Resources (Canada) Limited (formeþ Consolidated Cinola Mínes Ltd')o 'Queen Clnriotte Islands, B'C"

The Cinola Gold Project is a proposed open pit mine for the production of

approxima tely 6.4million tonnes of ore and waste rock per year' Approximately 68Vo

of the mine wastes are potentially acid producing'

Acid generating waste rock will be separated into two groups for separate disposal'

One quarter of the acid generating waste rock will be permanently submerged in the

tailings Ímpoundment while the remaining three quarters of the potentially acid

generating rock will be stored in a stockpile and backfilled to the open pit upon

completion of mining. The pit will subsequently be filled with water thus inundating the

backfilled rock.

5-44


Constituent Analyses

Nitrate '12.14 mg/litrel

Dissolved MetalsCuMoPblronAI

0.0190.08

0.0020.12o.44

mg/litremg/litremg/litremg/litremg/litre

Dissolved Oxygen (field) 9.6 mg/litre

Sp. Cond 1 540

pH - fieldpH - lab

7.5227.g2

Turbidity Low

Table 5-24

1984 Water QualitY DataPhoenix Ftooded/Decommiss¡oned Open Pit

]Exceeas drlnklng water standard.¿Estlmated

3 day detay between fleld and tab analyslr.

Table 5-25

Sulphate lon and D_isso.lved Gopper Concentrat¡onstn Prov¡dence Ëaker

11975: plant operat¡ng; last year talllngs pond offluent dlschargedto Provldence l¡ke.

Constituent, mg/litre Year

so -251671.2

19751

19844

Dissolved Cu 0.0120.001

1

1

973984

5-45


Environmental monitoring for water and groundwater quality, and acid generation will

be conducted throughout the construction, operation and reclamation phases of the

project.

5.2 Summary

The case studies which have been reviewed in this chapter demonstrate that some data

are available which usefully document some processes of concern to subaqueous

disposal; however, such illustrative studies are very few in number. The majority of

examples which have been described in the open and grey literature are often of limited

utility because they adopt too narrow a focus. For example, few conclusions about

diagenetic processes underwater can be drawn from comparative mineralogic studies in

the absence of interstitial water data. Howevcr, the studies which have been reviewcd

imply that, in general, oxidation of submerged tailings may be inhibited even in very

shallow waters, as suggested by the Mandy [.ake example. Similarly, no evidence for

oxidation and associated metal release was f<lund in Buttle [-ake, where sulphide-rich

tailings are overlain by an oxygenatecl and much deeper water column.

During the course of preparation of this chapter, no substantive evidence was

uncovered to support the notion that significant acid generation can occur under water.

This does not mean, however, that submergence categorically excludes oxidation; the

available information is too discrete and limited to permit such a generalization to be

drawn at this point, as is noted in the following chapter.

5-46

6.0 CONCLUSIONSANDRECOMMENDATIONS

The literature review presented in this report was conducted as a multidisciplinary

exercise to evaluate the current understancling of both acid generation and subaqueous

disposal practices and to establish a foundation for proceeding with further studies.

The additional studies, when completed, would provide site-specific information to fill

gaps in existing data and improve our understanding of the long-term behaviour of

active and abandonecl submerged mine waste deposits. Only then can a set of criteria

for environmentally safe methods of underwater disposal of reactive mine wastes be

established to permit confident decision-making on the suitability of aquatic discharge

strategies.

Much information was collectecl ancl reviewed in our study, a good deal of which is

summarized in the preceding sections. Basecl on the combination of theoretical and

case stu¿y data, a series of conclusions and recommendations follow which provide a

synopsis of the state-of-the-art in subaqueous disposal and demonstrate where further

field investigation should be completed to improve the level of understanding of the

physical, chemical and biological implications of storing reactive mine wastes

underwater.

6.1 Conclusions

Based on the theoretical review of AMD ancl case study investigations, the following

conclusions have been drawn. They consider both technical and non-technical issues,

and provide a foundation for the subsequent discussion of recommended further

studies. Individual conclusions are not mentioned in order of importance.

. AMD is a serious clisaclvantage for land-base<l tailing disposal, particularly in

wetter climates characterizing certain areas of British Columbia. The typical

oxidation-recluction (redox) reactions and galvanic interaction of commonly

occurring waste sulphicle minerals (pyrite, pyrrhotite) that occur in the presence

of oxygen, moisture and Tltíobacillus fenooxidaru (catalyst) can result in the

generation of highly aciclic water and concomitant leaching of trace metals from

mine wastes. By minimizing exposure to oxygen by storing reactive wastes

underwater, subaqueous disposal practices hold considerable promise for

6-1

a

a

CONCLUSIONS AND RECOMMENDATIONS

suppressing acid generation. Further study is required however to understand

more fully the conditions under which underwater disposal will be successful.

Tests to evaluate acid mine drainage and/or acid-generating potential of

lancl-based mine wastes appear to have very limited or marginal predictive

ability for wastes submergecl underwater (unless the body of water is very well

orygenated). Of the range of tests available, the kinetic shake flask test, which

gives an indication of the degree to which metals can cnter thc water column

when tailings are disposed into a lake, appears somewhat suitable for

underwater storage of reactive mine wastes as it best simulates actual field

conditions and has been shown to be accurate in its predictions.

Any databases which contain chemical andf or physical data for lakes currently

exist in a form too incipient to be of significant use for evaluating candidate

lakes as mine waste repositories. This reflects the relatively recent concern

about the long term impacts of subaqueous disposal.

Critical areas in the biological arena still remain that require more detailed

investigation, parricuiariy the complex processes of bioavaiiabiiity of metals in

lake-bottom sediments and bioaccumulation in the freshwater food chain. t¿ke

studies have not adequately evaluated post-depositional reactivity of submerged

sulphide-bearing tailings to determine if benthic effluxes of heavy metals are

present.

Macrobiological impacts must be examined under each of four categories:

turbidity, seclimentation, toxicity to freshwater biota and contamination of the

food web. Each category can represent significant macrobiological impacts; the

primary impacts of which are outlined as follows. Turbidity affects water

transparency and hence primary and secondary production as well as respiration,

feeding and other behaviour of water column organisms. Sedimentation results

in the smother¡ng of eggs and benthic organisms. Toxicity includes a wide range

of lethal, sub-lethal and behavioural impacts of trace metals or acid-generating

materials on freshwater biota. Contamination effccts consider the

bioaccumulation of trace metals through the food chain.

6-2

a


Various factors representing a wide array of processes, chemical reactions, and

site-specific conditions interact, perhaps synergistically, in determining whether

or not mine wastes deposited underwater have the potential for introducing

substances toxic to biological systems. Several factors worth considering include

the following:

the natural chemistry or chemical properties of the receiving

environment;

physicochemical c<lnclitions which may aid in reducing concentrations of

dissolved metal species through conversi<¡n into authigenic phases, i'e'

precipitation;

size of the water body and its capacity to tolerate a given amount of

effluent;

hydrochemical conditions that may increase heavy metal solubility;

the existence ancl frequency of flushing cycles in a lake;

the composition and pH of the mine wastes which can reflect a range of

metal recovery treatment processes - anywhere from state-of-the-art

selective flotation to those that have undergone some alteration, for

example, through cyaniclation or <lther hyclrometallurgical or roasting

pre-treatment; and

the existence and the types of aquatic life that inhabit the lake and their

migratory and feeding habits. Through certain mechanisms, biota have

the potential for transporting metals in the sediments or water column

into the food web.

The case stuclies reviewed, although consiclered representative, yield

inconclusive results. Data are generally sparse or superficial, only somewhat

relevant, reflect limited sampling, andfor have little predictive capacity to

permit improving our understanding of long-term phenomena associated with

subaqueous disposal. The data collected are usually for short-term

environmental impact assessments ancl not for detecting biogeochemical

6-3

a


processes or longer term trends. Water quality data are common but very little

sediment mineralory is available and virtualty no pore water chemistry.

Regular, comprehensive monitoring of water chemistry in flooded open-pits is

sparse at present. Such information, if available in significant quantity and

quality, could provide a database to assist in future predictions of the chemical

evaluation of drainage from abandoned pits.

6.2 Recommendations for Field lnvestigations

The preceding chapters demonstrate that the subaqueous disposal of reactive mine

wastes appears to be a highly promising alternative to land-based disposal, but that

critical gaps exist in the data which prevent establishing a set of criteria for effectively

evaluating alternative disposal strategies. Clearly, further field studies need to be

designed to evaluate post-depositional reactivity of submerged sulphide-bearing mine

wastes to determine if benthic effluxes of selected metals exist and to what extent they

are mitigated by the gradual cleposition of a natural sediment layer. Included must be

more detailed, site-specific investigations of impacts on the biological community native

to the receiving environment.

Although discussed individually below, it is most important that the geochemical and

biological components be conducted for each site investigation" This suggestion refiects

the necessity of increasing our knowledge of the factors which control metal release or

uptake by tailings and waste rock and the associated direct or indirect impacts on

aquatic organisms. It is only then that a complete understanding of the long-term

nature of submergecl mine wastes ancl their potential biological impacts can be fully

appreciated.

6.2.1 Geochemical Investigations

It is recommended that in future work, detailed studies be carried out on the

distribution of dissolved métals and metabolites in interstitial waters collected from

suites of cores raised from submerged tailings deposits in a number of lakes. Previous

work has demonstrated the utility of this approach in assessing the post-depositional

reactivity of tailings cleposits in such environments (e.g. Pedersen, 1983).

6-4

CONCLUS¡ONS AND RECOMMENDAT¡ONS

For comparative purposes, such studies should embrace a variety of deposits, including

unperturbed sediments and tailings which have contrasting mineralogies, and should

include assessment of alteration effects, and connate water and/or groundwater

chemistry in companion subaerial tailings deposits on land. In both types of deposits,

the work should include locations which are no longer receiving mine waste, as well as

active depositional regimes. It is strongly recommended that these investigations

inclucle chemical analyses of selected major and minor element concentrations in the

soli<l phases from which the pore waters are extracted, mineralogical characterization,

and measurements of the organic carbon concentration.

Where tailings or waste materials of the same composition have been discharged both

underwater and on land (as for example at Buttle and Benson l-akes), comparative

mineralogic studies of both facies should be carried out by X-ray diffraction to contrast

the extent and nature of alteration of minerals under on-land and submerged

conditions. The degree of such alteration is expected to be minor in the submerged

deposits, but this requires confirmation. Such comparisons with waste materials on

land, which are frequently heavily altered, have the potential to provide a particularly

illustrative suite of examples of the relative chemical diagenetic behaviour of tailings

and waste rock exposed to the atmosphere versus those under water.

To extract interstitial waters properly, consi<Jerable precautions must be taken to collect

undisturbed cores ancl to avoid oxidation and contamination of the samples. Such work

is, therefore, time and labour intensive and recognition of this requirement should be

made in future budget deliberations. Interpretation of the pore water chemistry will

require collection of supporting limnological information, including dissolved oxygen

and other hydrographic measurements.

Such studies, if carried out in sufficient detail, promise to deliver increased

understanding of processes governing the post-depositional release (or lack of release)

of dissolved metals to waters in the host basins, as well as permit calculations of the

metal effluxes or influxes from impacted lake floors.

Given the several stages of chemical evolution which are expected to characterize the

history of a submergecl deposit, it is suggestecl that the recommended studies

incorporate the following comp<lnents:

6-5

a


pore water work on a suite of representative cores should be performed one or

two years after the mines have commenced discharging in order to assess the

fluxes of metals out of or into the deposits. The purpose of this effort is to

monitor the post-depositional chemical behaviour of the deposits during the

active depositional stage. Ideally, characterization of the ambient chemistry of

interstitial waters should also be carried out in lake basins chosen for

sub-aqueous discharge before such disposal begins. Such pr'e-operational work

would permit <tefinition of metal releases for sediments which are known to

occur in some pristine lakes as a consequence <lf natural geochemical cycles.

For example, tailings discharge to Brucejack I-ake in northern B.C. is planned to

commence in 1990; it is recommended that a survey of the distribution of certain

dissolved metals in Brucejack l¿ke sedimentary porc waters should be madc

before tailings deposition commences. Such baseline data would prove to be

invaluable to subsequent comparisons with the chemistry of pore waters in the

tailings;

the same sampling strategy should be employed at two later stages:

1. within 12 months of cessation of diseharge, in order to assess the potential

effects of oxidation following the abrupt change in the nature and rate of

sedimentation; and

2" two to six years after cessation of discharge, to permit evaluation of the

potential "sealing" effect which is expected to be provided by the

accumulation of natural sediments on the surface of the waste deposit.

Potential targets in British Columbia for thc requircd detailed work include Buttle

I-ake, Benson [,ake, and Kootenay lake (cspecially the location where the 1960's spill

of high-gr ade lead-zinc concentrates occurred, where significant potential exists for

galvanic interaction between dissimilar sulphides in contact with each other, or in

proximity to the Bluebell Mine at Riondel). Outside British Columbia, Mandy [-ake,

Fox l¿ke or Anderson I-ake are suggested as suitable targets for study, given their

shallow depths, relatively small size, and the high sulphide contents in the submerged

tailings.

In addition to the foregoing, there are several British Columbia coastal marine

environments containing mine tailings that could be investigated, ideally by using the

6-6

CONCLUS¡ONS AND RECOMMENDATIONS

same stratery outlined above. This information could supplement that collected at the

freshwater lakes outlined above and aid in drawing conslusions as to the environmental

impacts of subaqueous disposal of míne wastes. Topics which require investigation

include: the nature and extent of remobilization of selected heavy metals from

sediments, chemical speciation and toxicity of different species, and the mechanisms

ancl extent of metat uptake by organisms. Sites in which such studies could most

effectively be carried out include:

. Rupert Inlet (Island Copper) - mine still active. Studies of early diagenesis

during the active depositional stage have been carried out previously (Pedersen,

1984, 1985).

. Howe Sound (which hosts tailings derived from the former Britannia Mine) -

investigation of pore water is currently undenvay by members of the

Department of Oceanography, U.B.C. Release of tailings to the fjord ceased in

t974.

Alice Arm (Kitsault/Amax Canada) - tailings deposition ceased November,

L982. Work on arsenic speciation in interstitial waters has been carried out at

Royal Roads Military College (Esquimalt), on Mo and metabolite chemistry by

I-osher (1985), and on Cu, Cd, Zn and Pb distributions by J.A.J. Thompson

(Ocean Chemistry, Institute of Ocean Sciences; unpublished data).

Anyox - site of B.C.'s (Canada's) first copper smelter; the local bay is thought to

contain mine, mill and smelter wastes.

a

lvith the exception of the Alice Arm, Rupert Inlet, Buttle I¿ke and Howe Sound

deposits, we suspect that very little monitoring or sampling (apart from water quality in

the overlying water columns) has been done at the great majority of abandoned sites of

submerged wastes. In many cases, it is clear that the "window of opportunity" has

passed in which it may have been possible to assess the initial post-abandonment

behaviour of waste materials in a number of water bodies. Such behaviour could then

have been compared with subsequent studies which would have permitted a long-term

chemical history to be derived. It will be possible however to assemble such diagenetic

histories for locations in which deposition is undenvay or has recently ceased, in

particular deposits such as those in Buttle l¿ke and to a lesser extent Benson Lake.

6-7


6.22 Limnological and Biological Investigations

Limnological and biological studies must be completed in association with geochemical

investigations to link the complex process of metals release from submerged mine

wastes to their uptake by aquatic organisms and bioaccumulation in thc food chain.

The purpose of the limnology/biolory program will be to dcscribc the lake in terms of

features that can be used to predict the impact of mine wastes deposited in similar

lakes. These measurements will allow investigators to calculate lake turnover and

residence time (flushing rate), determine circulation and mixing features, and evaluate

the potential for the waste material to be mobilized by meteorological or hydrological

events and the rate at which contaminants from the deposit would be dispersed' In

addition, the measurements will fill existing data gaps on metal transfcr from sediments

up through the aquatic food chain, a process which is currentþ poorly understood with

respect to submerged mine wastes.

Dissolved oxygen, pH, and redox potential are parameters which will influence the

reactivity of the sediments. Our assumption is that if the sediments and interstitial

waters are anoxic, then the solubility of metals will be minimized by the presence of

hydrogen sulfide. The research program would, ideally, investigate this hypothesis for

at least two, quite different lake types:

1. low productivity, high oxygen lakes which are high in alkalinity; and

Z. highly productive lakes which are anoxic or suboxíc for at least a portion of the

year.

Research conducted by one member of our study team suggests that tailings high in

calcium (e.g.lime-bearing tailings) may significantly reduce phosphate regeneration

from the sediments which, in turn, may influence primary production rates in thc lake'

This study would evaluate the sediment and water chemistry data collected to test this

hypothesis.

It is suggested that site-specific experiments on lacustrine biota be designed to establish

the impacts of heavy metals on both the infauna and epifauna. Organisms which reside

in the sediments (infauna) can have a significant influence on the degree of mixing that

occurs, which in turn may influence the mobilization of mctals. Mctal lcvcls within the

tissues of the infauna and epifauna may reflect metal uptakc rates and the potential for

6-8


transfer to, and bioaccumulation in other organisms' The mobility of fish may make

interpretation of metal levels in their tissues difficult to correlate to mine waste

disposal, particularly if historical or pre-discharge information is not available' Because

of the high interest in fish by both regutatory agencies and the public' however'

selection of one or two suitable species (e.g. based on low mobility, longer life-spans' or

higher trophic level feeders) is considered appropriate'

Based on a combination of theory and empirical case study data, initial attempts were

made at developing a decision model with which to evaluate the suitability of future

underwater waste disposal strategies. These efforts were confounded, however, by the

insufficient and/or unreliable data that exist which document conditions at existing

subaqueous disposal sites. This barrier to developing a systematic and comprehensive

method for evaruating discharge options reinforces the need for the research initiatives

outlined above. such research wilt directly assist in the development of a decision

model that will incorporate both physical theory and empirical' geochemical'

limnological and biological data in a critical path framework for evaluating the

environmental efficacy of subaqueous disposal of reactive mine wastes'

Examples of physical parameters to be included in the model include the extent and

profile of the mine wfrste deposit, lake bathymetry þarticularly near the disposal area)'

turnover patterns and flushing rates. Dissolved oxygen and pH, particularly near the

sediment/water interface, which along with analysis of dissolved metals and salts in the

lake water, represent some of the chemical parameters which must be considered'

Sediment chemistry, interstitial water chemistry, organic carbon content of sediments'

metal partitioning and speciation in sediments, the presence of oxidized minerals and

evidence of potassium scavenging in the sediments are ail pertinent considerations.

Biological/limnological parameters, as described earlier, must include lake

productivity, degree of bioturbation in sediments, and metal levels in representative

infauna, epifauna and fish species.

The decision model will incorporate the parameters of concern to subaqueous disposal

identified above and those refined through field study. once developed, the model will

have a very practicat application for both industry and government in the screening of

disposal alternatives based on a pragmatic fatal flaw approach that identifies key

potential problem areas with proposed discharge strategies' Overall' benefits from

developing and applying the decision model will accrue to industry in terms of

6-9

CONCLUS¡ONS AND RECOMMENDATIONS

effectively choosing methods to dispose of reactive mine wastes and to government

charged with the responsibility of ensuring wastes are disposed of in a manner

consistent with the protection of the natural environment.

6-10

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7-69

APPENDIX UA'

GLOSSARY OFTERMS

A-f

APPENDIX UA''

Advection

Allochthonous:

Allogenic:

Anthropogenic:

Authigenesis:

Autochthonous:

Autotrophic:

Benthic/Benthonic:

Benthos

Bioturbation:

Daphnia:

Within-lake unidirectional motion that transports matter in

lake systems; identity of transported substance unchanged.

Sedimentary rocks whose constituents have been

transported and deposited some distance from their origin.

Rock masses transported via tectonic forces.

Minerals brought into the lake by surface water, shorc

erosion, glacial transport, aeolean proccsses (cf.

endogenic).

Originated or caused by man.

Process by which minerals form in a sedimentary rock after

its deposition, i.e. formed where found (q.v. diagenesis).

Sedimentary rocks formed in place; bedrock masses that

remain in place in mountain belts.

Pertaining to organisms able to manufacture their own food

from inorganic substances.

Relating to, or occurring at the bottom of a body of water

and/or in ocean depths.

Aquatic organisms that live on or in the bottom of a body

of water.

Burrowing activity/mechanical disturbance caused by

benthic organisms or fauna; stirring of sediments by the

activity of burrowing benthonic organisms.

An important invertebrate freshwater crustacean, and onc

of the most sensitive to heavy metals.

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APPENDIX UA''

Diagenesls:

Endogenlc:

Epifauna:

Epilimnlon:

Euphausiids:

Euphotic:

Exuviae

Floodplain [¿ke:

Halocline:

FTU

Sum total of processes (physical, chemical and/or

biological) bringing about changes in sediment or

sedimentary rock after its deposition in water (q.v.

authigenesís).

Minerals originating from processes occurring within the

water column.

Organisms that inhabit the water column, remote from

sediments.

Upper zone of lake waters, generally well'mixed by surface

currents and wave action, and can be significantly warmer

than poorly-mixed deeper waters, i.e. hypolimnion.

An order of shrimp-like crustacean organisms, commonly

luminescent.

Relates to or constitutes the upper layers of a body of water

penetrated by sufficient light to permit growth of green

plants.

Cast-off skin, shell or covering, as in molting organisms,

animals.

L-ake subject to periodic silt influxes and nutrientladen

river water either by inundation or sometimes by

connections to river channels; receive highly variable

amounts of river water.

Water layer in which there is a large change in salinity with

depth.

Formazín Turbidity Units.

Refers to organic-rich fine-grained sediments found on

continental margins or the transitional zone between a

continental shelf and the deep sea.

Hemipelagic:

A-3

APPENDIX UAU

Heterotrophic:

Hypolimnion:

lnfauna:

lnstars

JTU:

Lacustr¡ne:

Lalces:

Requiring preformed organic compounds of nitrogen and

carbon for food; unable to manufacture food from

inorganic compounds.

Poorly-mixed deeper waters of a lake.

Organisms that inhabit/reside in sediments.

Discrete stages of growth or development

Jackson Turbidity Units.

Of, or pertaining to, or formed or growing in, or inhabiting

lakes.

Dimictic: A lake which circulates vertically twice a

year'

Dystrophic: Lake decaying to extinction.

Eutrophic Rich in dissolved nutrients (phosphate

and nitrate); often shallow and

seasonally deficient in orygen; denotes

degree to which lake has aged.

Holomictic: The whole lake volume undergoes, at

sometime, complete mixing - usually

taken to be approximately annually.

Meromictic: Part of the lake undergoes mixing;

bottom never mixes. Water mixing is

incomplete; non-circulating bottom

waters are isolated from circulating

upper waters, i.e. a PermanentlY

stratified lake.

Mesotrophic: Having a moderate amount of dissolved

A-4

nutrients.

APPENDIX UAU

Lysimeter:

Macrobenthos:

Mesocosm:

Metabolites:

Mixolimnetic:

Monimolimnetic:

Oligotrophic: Deficient in plant nutrients; has

abundant dissolved oxygen with no

marked stratification; i.e. sparse

productivity and sterile substrate.

Polymictic: Periodic, irregular, complete mixing, i.e.

tropical lakes.

A device for measuring the percolation of water through

soils and for determining the soluble constituents removed

in the drainage.

See Benthos; bottom organisms of larger size that cause or

contribute to bioturbation, q.v. microbenthos.

Intermediate to microcosm and macrocosm.

Compounds elaborated by the biological activity of

organisms andf or products of metabolism, i.e. NHo+ and

Pqt. Chemical products from biotransformation of

xenobiotica.

Mixing strata of water.

Non-mixing.

Structure or form of something. Basíc physical features of

lake, i.e. depth, atea, volume, perimeter, shoreline

development.

Morphology:

Mysids: A type of freshwater shrimp.

NTU: Nephelometric Turbidity Units.

Pelaglc: Pertaining to the open sea.

Phytoplankton: Microscopic plant forms of plankton.

Oxic-anoxic interface.Redoxcline:

A-5

APPENDIX IIA'I

Suspended Solids:

Taconite:

Thermocline:

Particulate matter in suspension, usually referring to

particles of >0.4 ¡.rm diameter.

A flint-like sedimentary rock, hard enough to cut glass; a

merchantable or non-merchantable ferruginous chert or

ferruginous slate comprising compact, siliceous rock

containing very finely disseminated iron oxide.

A transition zone or mixing layer in a thermally stratified

body of water that separates an upper' warmer' lighter,

oxygen-rich zonc from a lower, colder, denser zone which

may be oxygen-poor.

Trophic: Relating to nutrition or nutritional.

Isolated from river water; usually clear water bodies with

stable, well-vegetated banks, and bottoms that slope gently

from shore. Bottom substrates typically organic muck, but

shallow areas can be hard sandy material overlain by

smooth gravel and cobbles. Emergent and submergent

rooted vegetation is common. I¿kes are deepest in regions

of pronounced relief; may have bottom areas with abrupt

and considerable depth variations.

Upland Lake:

Vadose water: Subsurface water above the zone of saturation in the zone

of aeration.

Vadose Zone: Region/area with groundwater suspended or in circulation

above the water table.

Zoobenthos: Benthic fauna that lives on or in the lake bottom'

Zooplankton: Animal forms of plankton.

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